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Geochemistry of natural radionuclides in uranium-enrichedriver catchments

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Geochemistry of natural radionuclides in

uranium-enriched river catchments

A thesis submitted to the University of Manchester for the degree of Doctor of

Philosophy in the Faculty of Engineering and Physical Science

2013

Saif Eldin Mohammed Babiker Siddeeg

School of Chemistry

2

Table of Contents List of Figures ................................................................................................................. 5 List of Tables .................................................................................................................. 9

Abstract ........................................................................................................................ 11 Declaration ................................................................................................................... 12

Copyright Statement ..................................................................................................... 13 Dedication .................................................................................................................... 14

Acknowledgments ........................................................................................................ 15 Chapter One .................................................................................................................. 17

1 Introduction.......................................................................................................... 17

1.1 Research outline ................................................................................................... 17

1.2 Thesis structure .................................................................................................... 18

1.3 Review of U-series geochemistry ............................................................................ 18

1.3.1 Natural U-decay series ..................................................................................... 18 1.3.2 Uranium mineralogy......................................................................................... 26

1.4 Fractionation of U-series radionuclides in the surface environment ......................... 31 1.4.1 Chemical fractionation ..................................................................................... 32

1.4.2 Physical fractionation ....................................................................................... 34

1.5 U-series in surface waters ........................................................................................ 38

1.6 U- series fractionation in river waters ...................................................................... 39 1.6.1 Weathering effect ............................................................................................. 39

1.6.2 Fractionation processes during river transportation ........................................... 40

1.7 Natural analogues .................................................................................................... 42

1.7.1 Natural analogues in the UK ............................................................................. 43

References .................................................................................................................... 47

Chapter Two ................................................................................................................. 56 Development of radium separation ................................................................................ 57

2 Introduction.......................................................................................................... 57

2.1 Radium at South Terras mine ............................................................................... 57

2.2 Techniques for radium measurement .................................................................... 58

2.2.1 Radiometric techniques ............................................................................... 59 2.2.2 Atom Counting Techniques ......................................................................... 62

2.3 Methodology ........................................................................................................ 63 2.3.1 Preparation of radium test solution .............................................................. 63

2.3.2 Radiochemical separation testing................................................................. 64 2.3.3 Alpha source preparation ............................................................................. 65

2.3.4 Sample measurement ................................................................................... 65

3

2.4 Results and discussion .......................................................................................... 66

2.5 Conclusions and recommendations ....................................................................... 70

References .................................................................................................................... 71

Chapter Three ............................................................................................................... 74 Dispersion of U-series radionuclides in stream sediments from Edale Valley, UK ......... 75

Abstract .................................................................................................................... 75 3 Introduction.......................................................................................................... 76

3.1 Naturally occurring uranium................................................................................. 76

3.2 Fractionation of 238

U-series .................................................................................. 77

3.3 Objectives of the study ......................................................................................... 78

3.4 Materials and methods.......................................................................................... 79

3.4.1 The study area ............................................................................................. 79 3.4.2 Sampling and sample pretreatment .............................................................. 80

3.4.3 Mineralogy of the samples .......................................................................... 80 3.4.4 Radiochemical characterisation ................................................................... 80

3.4.5 Total radium ................................................................................................ 82 3.4.6 Radium separation ....................................................................................... 83

3.4.7 Quality control ............................................................................................ 84

3.5 Results and discussion .......................................................................................... 85

3.5.1 Characterisation of the stream sediments ..................................................... 85 3.5.2 238

U, 234

U, 230

Th and 226

Ra contents of the sediments ................................... 85

3.5.3 Fractionation of the radionuclides ................................................................ 86 3.5.4 234

U/238

U and 230

Th/238

U isotopic ratio diagram ........................................... 88

3.5.5 Hierarchical cluster analysis ........................................................................ 88

3.6 Conclusions.......................................................................................................... 93

Acknowledgements ....................................................................................................... 93

References .................................................................................................................... 94

Chapter Four ................................................................................................................115 Geochemical characterisation of uranium and radium in sediments near an abandoned

uranium mine, Cornwall, UK .......................................................................................116

Abstract .......................................................................................................................116

4 Introduction.........................................................................................................117

4.1 The study area and sampling ...............................................................................118

4.2 Methodology .......................................................................................................119 4.2.1 Physicochemical analysis of water..............................................................119

4.2.2 Physicochemical properties of sediments ....................................................120

4

4.2.3 Radioactivity content in sediments .............................................................121

4.2.4 Total radium in sediments ..........................................................................122 4.2.5 Sediment characterisation ...........................................................................126

4.3 Results and discussion .........................................................................................128 4.3.1 Physico-chemical properties of stream waters.............................................128

4.3.2 Physico-chemical properties of sediments ..................................................128 4.3.3 Radiochemical characterisation of sediments ..............................................129

4.3.4 Sequential chemical extraction results ........................................................130 4.3.5 Radionuclide and stable element fractionation ............................................131

4.3.6 Uranium isotopic ratios in sequential extraction fractions ...........................133 4.3.7 Characterisation of sediments using spectroscopic methods ........................134

4.4 Conclusions.........................................................................................................135

Acknowledgements ......................................................................................................136

References ...................................................................................................................138 Chapter Five ................................................................................................................159

Conclusions and Recommendations .............................................................................159

Future work .................................................................................................................161

Appendix .....................................................................................................................162 Methods and experimental techniques ..........................................................................162

A1 Areas of the study ..................................................................................................162 A1.1 Edale sampling ................................................................................................162

A1.2 Cornwall sampling ..........................................................................................162

A2 Sediment analysis ...................................................................................................166

A2.1 Loss on ignition ...............................................................................................166 A2.2 Radiometric techniques ...................................................................................166

A2.2.2.1 Sample preparation ........................................................................................177 A2.3 Mineral analysis techniques .............................................................................178

A3 Water analysis ........................................................................................................184 A3.1 Physicochemical properties .............................................................................184

A3.2 Ion chromatography ........................................................................................185

References ...................................................................................................................187

5

List of Figures

Figure 1.1 U-decay series with long and intermediate half-life members in square shapes

(Lyczkowski, 1982) ......................................................................................................... 20

Figure 1.2 Species of uranium in a system containing [UO22+

] = 10 µM and [CO32-

] = 10

mM as a function of pH at 25 ⁰ C, using MEDUSA* ........................................................ 23

Figure 1.3 Uraninite (dark crystals) in brecciated matrix (www.webmineral.com) ........... 28

Figure 1.4 Black uraninite on yellow uranophane (www. webmineral.com) .................... 28

Figure 1.5 The first part of the 238

U-series ....................................................................... 32

Figure 1.6 A diagram summarising alpha recoil effect, and chemical and physical

fractionation of U-series nuclides as a function of time (Dosseto et al., 2008) .................. 37

Figure 1.7 Edale Valley sediments 234

U/238

U vs 230

Th/238

U diagram as an example of a

complex U-series disequilibrium, with the comlex zone in grey colour ............................ 41

Figure 3.1 Edale Valley, Derbyshire and the sampling points .........................................106

Figure 3.2 234

U/238

U activity ratios from total dissolution analyses of sediments from Edale

Valley .............................................................................................................................107

Figure 3.3 234

U/238

U activity ratios from aqua regia leaching analyses of sediments from

Edale Valley ...................................................................................................................108

6

Figure 3.4 230

Th/238

U activity ratios from total dissolution analyses of sediments from

Edale Valley ...................................................................................................................109

Figure 3.5 230

Th/238

U activity ratios from aqua regia leaching of sediments from Edale

Valley .............................................................................................................................110

Figure 3.6 234

U/238

U vs 230

Th/238

U diagram for total dissolution analyses of sediments from

Edale Valley (Grey colour represents complex zones) .....................................................111

Figure 3.7 234

U/238

U vs 230

Th/238

U diagram for aqua regia leaching of sediments from

Edale Valley (Grey colour represents complex zones) .....................................................112

Figure 3.8 Dendrogram illustrating cluster analysis, from total dissolution data, of

sediments from Edale Valley based on five variables: [238

U], [234

U], [230

Th], [226

Ra] and

loss on ignition................................................................................................................113

Figure 3.9 Dendrogram illustrating cluster analysis, from aqua regia leaching, of


U], [234

U], [230

Th], [226

Ra] and

loss on ignition................................................................................................................114

Figure 4.1 Cornwall map showing the sampling points along the river Fal .....................147

Figure 4.2 Extraction profile of uranium as a percentage of the sum of five fractions in S3

and S7 .............................................................................................................................148

Figure 4.3 Extraction profile of radium as a percentage of the sum of five fractions in S3

and S7 .............................................................................................................................148

7

Figure 4.4 Extraction profile of calcium as a percentage of the sum of five fractions in S3

and S7 .............................................................................................................................149

Figure 4.5 Extraction profile of manganese as a percentage of the sum of five fractions in

S3 and S7 ........................................................................................................................149

Figure 4.6 Extraction profile of iron as a percentage of the sum of five fractions in S3 and

S7 ...................................................................................................................................150

Figure 4.7 Extraction profile of arsenic as a percentage of the sum of five fractions in S3

and S7 .............................................................................................................................150

Figure 4.8 Extraction profile of titanium as a percentage of the sum of five fractions in S3

and S7 .............................................................................................................................151

Figure 4.9 Extraction profile of barium as a percentage of the sum of five fractions in S3

and S7 .............................................................................................................................151

Figure 4.10 234

U/238

U activity ratios in the sequential extraction fractions of S3 .............152

Figure 4.11 234

U/238

U activity ratios in the sequential extraction fractions of S7 .............153

Figure 4.12 Scanning electron microscope (SEM) results showing backscattered electron

(BSE) image (A) and electron-dispersive spectroscopy (EXD) spectrum (B) of the bulk

minerals of S3. The bright area is an indication of a presence of high atomic number

element. ..........................................................................................................................154

8

Figure 4.13 Scanning electron microscope (SEM) results showing secondary electron (SE)

image (A) and electron-dispersive spectroscopy (EXD) spectrum (B) of the bulk minerals

of S3. ..............................................................................................................................155


image (A) and electron-dispersive spectroscopy (EXD) spectrum (B) of a single grain

isolated from the heavy minerals separated by heavy liquid from the richest U-sample (S7).

U association with P, Th and Ca (from the EDX analysis) has been identified. ................156


image (A) and electron-dispersive spectroscopy (EXD) spectrum (B) of the heavy minerals

separated by heavy liquid fractionation of S7. U associates with the aluminosilicates has

been identifird . ...............................................................................................................157

Figure 4.16 Backscattered electron (BSE) image (top) and X-ray maps of elements (Mn, K,

Fe, As, Ca and U) from electron microprobe analysis (EMPA) of the heavy minerals

separated by heavy liquid of S7 .......................................................................................158

Figure A1 A single grain identifying possible fractions released in sequential extraction

(Kaplan and Serkiz, 2001) ...............................................................................................172

Figure A2 Schematic diagram illustrating one separation step in sequential extraction

(Schultz et al., 1998b) .....................................................................................................173

Figure A3 Basic components of an ICP-MS. Adapted from Thomas (2008) ...................184

9

List of Tables

Table 1.1 Natural uranium isotopes with their half-lives and natural abundance .............. 21

Table 1.2 The origin, decay mode, half-life and radiation energy of the naturally occurring

radium isotopes ................................................................................................................ 25


radium isotopes ................................................................................................................ 58

Table 2.2 Optimising radium separation conditions in spiked nitric acid test samples and

226Ra chemical recovery ................................................................................................... 67

Table 3.1 Edale sediment sample coordinates, loss on ignition and mineralogy ............... 98

Table 3.2 The measured, the recommended and the leached values of 226

Ra and 238

U in

IAEA-314 stream sediment reference material ................................................................. 99

Table 3.3 Activity concentrations (Bq.kg-1

dry weight) of the total 238

U, 234

U, 230

Th, 226

Ra

and 234

U/238

U, 230

Th/238

U, 226

Ra/238

U activity ratios of sediments from the Edale Valley (±

1σ counting statistics uncertainties) .................................................................................100


dry weight) of the leached 238

U, 234

U, 230

Th,

226Ra and

234U/

238U,

230Th/

238U,

226Ra/

238U activity ratios of sediments from the Edale

Valley (± 1σ counting statistics uncertainties) .................................................................101

Table 3.5 Average activity concentrations (Bq.kg-1

dry weight) of Edale sediments (total

dissolution) and loss on ignition (wt.%) of the hierarchical cluster analysis (S.D. = standard

deviation). .......................................................................................................................102

10


dry weight) of Edale sediments

(leached) and loss on ignition (wt.%) of the hierarchical cluster analysis (S.D. = standard

deviation). .......................................................................................................................104

Table 4.1 Summary of the sequential extraction method applied for radionuclides and

stable elements from Cornwall sediments (sample/reagent ratio is 1.0 g/ 15 .0 mL) .........124

Table 4.2 Physiochemical properties, anions of water samples collected from the River Fal

and side streams in Cornwall and the coordinates of the sampling points ........................142


Ra and 238

U in

IAEA-314 stream sediment reference material ................................................................143

Table 4.4 Concentrations of cations in mg/L (g/L for Cu, As, Pb and U) in the filtered

water samples (<0.22 m) collected from the River Fal and side streams in Cornwall .....144

Table 4.5 Mineralogical composition from XRD and loss on ignition of sediments

collected from the River Fal and side streams in Cornwall ..............................................145

Table 4.6 U-isotopes and Ra activity concentrations (Bq kg-1

dry weight) and isotopic

ratios in 20 sediment samples collected from locations around the River Fal and side

streams in Cornwall (± 1σ counting statistics uncertainties) ............................................146

Table A1 Sample locations from the River Noe in the Edale Valley, the Peak District ....164

Table A2 Sample locations from the valley of the River Fal, Cornwall ...........................165

Table A3 Sequential extraction steps ..............................................................................171

11

Abstract The University of Manchester

Saifeldin Mohammed Babiker Siddeeg

Doctor of Philosophy

Geochemistry of natural radionuclides in uranium-enriched river catchments

2013

Radionuclides from natural U-series in sediments from two river catchments in the UK

have been studied. The aim was to gain insight into the behaviour of 238

U, 234

U, 230

Th and 226

Ra in real natural systems enriched in uranium. A radiochemical method for radium

separation followed by alpha spectrometric measurement has been developed. The method

allowed use of 225

Ra, in equilibrium with the parent 229

Th, as a yield determinant, and has

been applied in 226

Ra concentrations measurements in the selected areas of study.

U-series progeny, 238

U, 234

U, 230

Th and 226

Ra, in totally dissolved sediments from the valley

of the River Noe and the fraction leached by aqua regia, have been measured. Total

sediment contents ranged from 9 ± 2 to 184 ± 8 Bq.kg-1

for uranium, 9 ± 3 to 200 ± 13

Bq.kg-1

for thorium and 18 ± 1 to 179 ± 8 Bq.kg-1

for radium. The activity concentrations in

the leached fractions, compared with the total, were 46% for uranium, 54% for thorium and

56% for radium, on average. The radionuclides showed extensive disequilibrium and this

suggested a complex leaching/accumulation of uranium as well as an impact of organic

matter and secondary minerals.

Uranium and radium have been geochemically characterised in sediments from near the

South Terras abandoned uranium mine, Cornwall. Background activity concentration levels

of uranium in sediments ranged from 13 ± 3 to 290 ± 14 Bq.kg-1

, with radium from 42 ± 4

to 424 ± 23 Bq.kg-1

. Elevated concentrations of uranium and radium were measured in two

samples, S3 with 1820 ± 36 Bq.kg-1

for uranium and 940 ± 53 Bq.kg-1

for radium; and S7

with 4350 ± 53 Bq.kg-1

for uranium and 1765 ± 48 Bq.kg-1

for radium. Sequential chemical

extraction for the two samples revealed that both uranium and radium were associated with

organic and carbonate fractions, with 25 % of the uranium in the resistant phase of S7. 234

U/238

U activity ratios of the sequential extraction fractions showed different trends in the

sediments, and this was linked to the impact of organic matter and/or exchange between

water and sediment. Uranium-bearing minerals in association with potassium, calcium,

iron, manganese and arsenic have been identified in these sediments.

12

Declaration

No portion of the work referred to in the thesis has been submitted in support of an

application for another degree or qualification of this or any other university or other

institute of learning.

13

Copyright Statement

i. The author of this thesis (including any appendices and/or schedules to this

thesis) owns certain copyright or related rights in it (the “Copyright”) and s/he

has given The University of Manchester certain rights to use such Copyright,

including for administrative purposes.

ii. Copies of this thesis, either in full or in extracts and whether in hard or

electronic copy, may be made only in accordance with the Copyright, Designs

and Patents Act 1988 (as amended) and regulations issued under it or, where

appropriate, in accordance with licensing agreements which the University has

from time to time. This page must form part of any such copies made.

iii. The ownership of certain Copyright, patents, designs, trade marks and other

intellectual property (the “Intellectual Property”) and any reproductions of

copyright works in the thesis, for example graphs and tables (“Reproductions”),

which may be described in this thesis, may not be owned by the author and may

be owned by third parties. Such Intellectual Property and Reproductions cannot

and must not be made available for use without the prior written permission of

the owner(s) of the relevant Intellectual Property and/or Reproductions.

iv. Further information on the conditions under which disclosure, publication and

commercialisation of this thesis, the Copyright and any Intellectual Property

and/or Reproductions described in it may take place is available in the

University IP policy (see http://documents.manchester.ac.uk/

DocuInfo.aspx?DocID=487), in any relevant Thesis restriction declarations

deposited in the University Library, The University Library’s regulations (see

http://www.manchester.ac.uk/library/aboutus/regulations) and in The

University’s policy on Presentation of Theses.

http://documents.manchester.ac.uk/

14

Dedication

To my father and my mother,

To Khadiga and Lujain,

To my sisters and my brothers

15

Acknowledgments

I thank Allah almighty for his guidance, care and for many privileges known to me in

several and very specific ways.

I appreciate the generous financial support of my sponsor, Islamic Development Bank

(IDB), Jeddah, Saudi Arabia.

I wish to express my sincerest grateful to my supervisor, Professor Francis Livens for his

encouragement and support from the very early days and his valuable comments. Thanks

for the trips to South Terras, Cornwall and to Edale during snowy weather! I will not forget

to say thanks, on behalf of Lujain, for the one to Monte Verita, too!

Many thanks to Dr Nick Bryan, my co-supervisor, for his advice and guidance throughout

my time at CRR, and for joining the sampling team, digging in mud, even in Cornwall!

Special thanks to staff of the Geochemistry Analytical Unit and the Minerals Analysis

Facility (SEAES), Mr Paul Lythgoe, Mr Alastair Bewsher, Mrs Cath Daveis, Dr John

Charnock and Dr John Waters. All were helpful in training me during analysis of my

samples and in offering me space in their labs, even at extremely busy time.

I had a wonderful time with many people, past and present, in the Centre for

Radiochemistry Research, especially with Monday night football group. Thanks to

Mustafa, Ragiab, Hamza, Sean, Carlos, Nigel, Drew, Raj, Rick, James, Dan, Kurt, Ryan,

Simon, Tamara, Kate, Lucy, Daisy, Debbie, Gotfried, Jen, Maddie, Tony, Nick S, Ally,

Nick MW, Katie, Adam, Mike, Gareth, Clint, Sarah, Steph and Louis.

16

Socially, I had a delightful time in the UK with many people! Warm thanks to all those

who provided emotional support, especially when I was walking with crutches for eight

weeks! My thanks extend to my friends and colleagues at SAEC for support.

Ultimate thanks to my father, mother, sisters and brothers for continuous encouragement

when I am with them or away from them. I am pretty sure that when things seems

ambiguous, your sincere Duaa lights my way again.

Finally, much is owed to my family, with profound thanks to my wife Khadiga, for her

sustain love, understanding, patience and taking care of Lujain and me for more than three

years. Your support is highly appreciated. My daughter Lujain! I know one day we will

remember all the incredible things you have done in Manchester, and I think you will

recognize most of them!

17

Chapter One

1 Introduction

1.1 Research outline

The need to understand the processes operating within a natural or man-made geological

system have ensured that the natural decay series have a broad range of applications.

Geochemical differences between members of the naturally occurring 232

Th, 238

U and 235

U

decay series can lead to fractionation of the different elements in natural systems. The U-

series radionuclides display greater geochemical diversity than those in the 232

Th series,

and the higher concentrations of 238

U in nature relative to 235

U mean that the 238

U series is

of greatest interest. Within the 238

U series, 226

Ra (1600 year), 230

Th (7.5 x 104 year),

234U

(2.5 x 105 year) and

238U (4.5x 10

9 year) all have intermediate or long half-lives, which

render them suitable for use as geochemical tracers. This can be of enormous value in

researching the behaviour of radionuclides in radioactive waste repositories.

The main objectives of this project were:

i. To develop a method for radiochemical analysis of 226

Ra in environmental

samples using 225

Ra as a tracer.

ii. To use the 238

U series to explore radionuclide transport processes in the

uranium-enriched systems, the valley of the River Noe, Edale, Derbyshire; and

iii. To use the area around the former uranium mine site at South Terras, Cornwall

as a natural analogue for geochemical characterisation of uranium and radium.

18

1.2 Thesis structure

The thesis has been prepared in alternative format, and three of the chapters will be

submitted to relevant scientific journals for publication. The first chapter outlines the

background of U-series geochemistry. It provides insight into natural U-decay series,

chemical and physical fractionation among parents and daughters within the series, the

factors affecting the fractionation of U-series radionuclides in river waters and ends with

natural analogues as locations where the geochemistry of the U-series can be investigated.

The second chapter describes the development of a radiochemical separation of 226

Ra, as an

important radionuclide from uranium mining activities. The third chapter is an attempt to

understand U-series disequilibrium in an area known to contain uranium phosphate

minerals, Edale, Derbyshire. The fourth chapter explores geochemical associations of

uranium and radium from an abandoned uranium mine, South Terras, Cornwall, with

various parameters such as organic matter, minerals and distribution within sediments. The

final chapter draws overall conclusions to the project and suggests recommendations for

future work.

1.3 Review of U-series geochemistry

1.3.1 Natural U-decay series

The study of radioactivity was initiated in the late 19th century by the discovery by

Becquerel in 1896 of the natural radioactivity of uranium. In 1898, the Curies developed

the separation of other radioactive elements (e.g. polonium from uranium). This was

followed by separation of other naturally occurring radioactive elements from minerals

(Krishnaswami and Cochran, 2008). The radioactive elements were initially classified

based on the decay type (e.g. alpha and beta), and the existence of three natural decay series

19

(uranium, thorium and actinium). In particular, the U-decay series (Fig. 1.1), with its

radioactive isotopes of eight elements (U, Pa, Th, Ra, Rn, Po, Bi and Pb), has many

applications in geochemistry and earth sciences, which have been highlighted recently

(Chabaux et al., 2008; Krishnaswami and Cochran, 2008; Pekala et al., 2010). For instance,

in geochemistry, the mobility of uranium and its daughters in different geological

environments (e.g. surface and ground waters) has been investigated (Kronfeld et al.,

2004); while in Earth sciences, radioactivity is a fundamental tool in dating events in Earth

history (Chabaux and Bourdon, 2006; Ivanovich and Harmon, 1992). The main parameters

directing these applications are the diverse geochemical properties of the different U-series

radionuclides, in addition to the wide range of half-lives of the chain members

(Krishnaswami and Cochran, 2008). The interest in this study is in the longer and

intermediate half-life radionuclides in the natural U-decay series; namely, the parent 238

U

and the daughters, 234

U, 230

Th and 226

Ra.

20

Figure 1.1 U-decay series with long and intermediate half-life members in square shapes

(Lyczkowski, 1982)

1.3.1.1 Uranium

Uranium (atomic number 92) is an actinide element that was discovered in 1798 by

Klaproth, after dissolving pitchblende (U3O8) in nitric acid. Uranium has three naturally

occurring isotopes, 238

U, 235

U and 234

U, with different half-lives and natural abundances

(Table 1.1). Two of the uranium isotopes are members of the U-decay series (the parent

238U and the daughter

234U), while the isotope

235U is the parent of the Ac-decay series

(Ivanovich, 1992). Historically, uranium was used as a colouring agent for glass; however,

following the discovery of radioactivity, it was regarded as the most important radioactive

21

element. It is involved in nuclear energy production for both peaceful and non-peaceful

applications.

The concentration of uranium in the Earth’s crust is in the range of 2-4 ppm, and it is

known to be in higher concentrations in particular types of rocks (e.g. granitic rocks,

phosphates and black shales). Uranium has a large number of known minerals (~ 200), with

uraninite (UO2) and pitchblende (U3O8) as the most common primary minerals, though

many other secondary minerals (e.g. autunite, torbernite and uranophane) are known. In

natural waters, uranium concentrations display a wide range; for example, it is within the

range of 2-4 ppb in sea water and 0.1 ppb to 1 ppm in fresh water (Lehto and Hou, 2010a).

Table 1.1 Natural uranium isotopes with their half-lives and natural abundance

Isotope Half-life (years) Natural abundance (atom %)

238U 4.47 x 10

9 99.28

235U 7.04 x 10

8 0.72

234U 2.46 x 10

5 0.0058

In aqueous systems, uranium occurs in various oxidation states; trivalent, tetravalent,

pentavalent and hexavalent. Among these, the tetravalent and the hexavalent oxidation

states are dominant in environmental conditions. In oxidising environments, such as surface

water, hexavalent uranium is more stable, generally forming uranyl species (Lehto and

Hou, 2010a). In reducing environments, such as deep ground water, tetravalent uranium is

dominant. Generally, uranium in the lower oxidation state U(IV) is less soluble compared

with the higher oxidation state U(VI). The behaviour of uranium in aqueous media is

complex, with many parameters affecting its chemical/physical forms (speciation) and its

22

transport (Fig. 1.2). The important factors include: organic matter, secondary minerals,

physicochemical properties and complexing agents (Lehto and Hou, 2010a).

23

Figure 1.2 Species of uranium in a system containing [UO22+

] = 10 µM and [CO32-

] = 10

mM as a function of pH at 25 ⁰ C, using MEDUSA*

* Making equilibrium diagrams using sophisticated algorithm

24

1.3.1.2 Thorium

Thorium is an actinide element (atomic number 90), with six naturally occurring isotopes.

The thorium isotope, 232

Th (half-life = 1.4 x 1010

y), is a parent of the Th-decay series. The

average concentration of thorium in the Earth’s crust is 3-4 times that of 238

U. The principal

thorium minerals are monazite (a Th-containing lanthanide phosphate) and thorite

(ThSiO4). Two of the thorium isotopes are members of the U-decay series; 234

Th (half-life

= 24 d) and 230

Th (half-life = 75 x 103 y). The latter is an alpha emitter and it is important in

studying U-series disequilibrium (Lehto and Hou, 2010a).

Thorium occurs in the tetravalent oxidation state in solutions, and for this reason, it is

widely used as an analogue to tetravalent transuranic elements (e.g. Pu4+

and Np4+

). Similar

to uranium, speciation in natural systems depends on many factors. However, thorium

chemistry in solution is simple compared with uranium, since it occurs in only one

oxidation state (Lehto and Hou, 2010a). For example, tetravalent thorium ions are reported

to occur only in acidic media with pH < 2-3 (Plater et al., 1992). In the absence of

complexing agents, polymeric hydrolysis products of thorium are likely to be present in

fresh water. These thorium hydroxide complexes may adsorb onto surfaces of clay minerals

and humic acids, enhancing the removal of thorium from the solution. However, in natural

waters, preferential complexation with organic ligands is more dominant than polymeric

hydrolysis (Lehto and Hou, 2010a).

1.3.1.3 Radium

Radium, with atomic number 86, belongs to Group II of the Periodic Table. Its discovery in

1898 is associated with the Curie family, when early radiochemical separation was

employed to isolate the element radium from pitchblende (Jia and Jia, 2012). Radium,

25

together with calcium, barium and strontium, are sometimes called the alkaline earth

metals. Although the light Group II elements, beryllium and magnesium, are quite different,

radium and barium exhibit extremely similar chemistry. This makes the separation of

radium from Group II elements, particularly Ba, difficult (Lehto and Hou, 2010b). The

occurrence of radium in nature results from the U and Th decay series.

The natural decay series of the three radionuclides (U, Th and Ac) are the main source of

the four radium isotopes (223

Ra, 224

Ra, 226

Ra and 228

Ra) in the environment (Vasile et al.,

2010). Table 1.2 gives a summary of the origin, decay mode, half-life and the energy of the

natural radium isotopes. Among radium isotopes, 226

Ra is the most influential radionuclide

of the radium isotopes due to its potential hazard to the environment, even at low

concentrations (Aguado et al., 2008)


radium isotopes

Isotope Decay series Decay mode Half-life Energy (MeV)

223Ra Actinium alpha 11 d 5.72 (51.6%) 5.61 (25.2 %)

224Ra Thorium alpha 3.6 d 5.69 (94.9 %) 5.45 (5.1 %)

226Ra Uranium alpha 1.6 x 10

3 y 4.78 (94.5 %) 4.60 (5.6 %)

228Ra Thorium beta 5.8 y 0.046 Emax

26

1.3.2 Uranium mineralogy

The metal uranium has attracted global attention due to its unique characteristics, including

its radioactive properties. Since the early 19th

century, the primary focus was on uranium

exploration and exploitation. More recently, there have been increased concerns over both

the environmental impact and the radiological hazard of past and present U-mining

activities (Suresh et al., 2011). Therefore, studying the geochemical behaviour of uranium,

including its reactive transport chemistry, both experimentally and using geochemical

models, is necessary (Brown et al., 2010; Frostick et al., 2011). As a result of the different

chemical conditions of formation, uranium minerals display a wide range of structural and

chemical variability and around 200 uranium minerals are known (section 1.3.1.1).

Research into uranium behaviour in the environment contributes to improved understanding

of the development of different uranium deposits. It can also provide insight into possible

migration and retardation mechanisms affecting uranium and its decay products. Moreover,

understanding the geochemistry of uranium can help to predict, at least partially, the long-

term behaviour of spent nuclear fuel and other uranium-bearing waste materials (Strok and

Smodis, 2010). However, the immediate concern over uranium contamination is focused on

uranium processing facilities and operating mines. Sites contaminated with uranium have

been investigated, applying a range of different techniques to address the fate of uranium

and its decay product radionuclides in these environments (Brown et al., 2010; Um et al.,

2010).

1.3.2.1 Uranium minerals (Primary and Secondary)

Uranium deposits have been studied in detail and their geological settings can be classified

broadly into 14 groups (Finch and Murakami, 1999), including: sandstone; surficial;

27

volcanic; coal; lignite; phosphorite; metamorphic and black shales. Uranium tends to

precipitate in reducing environments, such as shales, or sandstone rich in organic matter or

iron sulphides, or in phosphate-rich sediments. Since uranium is also concentrated in

organic matter, lignite and coal comprise another group. Uranium is preferentially

partitioned into low temperature melts, which leads to concentration and production of

uranium-rich granites (Plant et al., 1999).

Commonly, uranium minerals are classified as being either primary or secondary. The first

group is uranium oxides with various U/O ratios, such as uraninite (UO2) and pitchblende

(U3O8), which are characterised by black and dark colours (Fig. 1.3). The second group

forms from alteration of primary minerals by chemical and physical processes (e.g.

hydration and oxidation). These are often U(VI) minerals formed by transport away from

the parent primary mineral and subsequent precipitation. These are characterised by light

yellow, green and red colours, exemplified by minerals such as uranophane

(Ca(UO2)2(SiO3OH)2·5H2O; yellow), as in Fig. (1.4), and torbernite

(Cu(UO2)2(PO4)2·12H2O; green).

28

Figure 1.3 Uraninite (dark crystals) in brecciated matrix (www.webmineral.com)

Figure 1.4 Black uraninite on yellow uranophane (www. webmineral.com)

http://webmineral.com/

29

1.3.2.2 Uranium mining

Uranium mining is an essential step of the front-end processes of the nuclear fuel cycle.

The process involves the chemical extraction of uranium from uranium-bearing minerals in

rocks. Once the uranium has been leached, it is precipitated as yellow cake using alkaline

reagents, such as ammonia and magnesium oxides (Gupta et al., 2004). Commonly, mining

of uranium ore is achieved by applying two main techniques; open-pit mining and

underground mining. When uranium lodes lie close to surface, the overburden is often

removed to access the ore. However, where the ore is deeper, drilling is required to

establish shafts to reach it (Hore-Lacy and Cutler, 2004). In-situ leaching (ISL) has become

more widely used in uranium mining in more recent decades (Mudd, 2002). In ISL, acidic

or alkaline solution, depending on the ore type, is injected and circulated underground to

leach uranium. The uranium solution is pumped back to a surface treatment plant, without

significant underground disturbance.

Uranium mining has an environmental impact both during and following mining activities

(Carvalho, 2010); this include raising radioactivity levels in the nearby environment (e.g.

soil, surface and ground water) by exposing waste rocks and tailings to weathering

processes. In underground mines, limited amounts of waste are produced compared with

open-pit mining, the environmental impact associated with this type is less significant.

However, ISL can lead to leaching and mobilisation of toxic metals, with the associated

risks to aquifers and water supplies (IAEA, 2011).

Mineral processing and the extraction of uranium ore result in spoil heaps and mill tailings

(Landa, 2004). Spoil heaps are comprised of the undesired material generated during the

exploration and extraction of the ore. Uranium mill tailings are the sand-like remains left

30

once the uranium has been leached out of the ore. Uranium mill tailings retain most of the

radioactivity since they hold most of the decay products present in the original ore. In a

low-grade ore (< 0.1 % U3O8), the percentage of the waste can reach > 90% of the ore

processed (Tripathi et al., 2008). The waste represents a heterogeneous mixture of minerals

with different grain sizes and radionuclides, making their separation challenging

(Lottermoser and Ashley, 2006; Wiles, 1983). The resulting radiological hazards arise

from: (i) the short-lived decay products of radium isotopes; and (ii) the similarity in

chemistry of radium with alkaline earth metals (Ca, Sr and Ba), so radium can easily

replace them in biological systems.

Inappropriate management of uranium mining and ore processing wastes is a potential

source of radioactive contamination. Mobility of solid particles and aqueous radionuclides

from the waste have the potential to increase contamination of the surrounding environment

(Tripathi et al., 2008). Dispersion of radionuclides toward surface water is likely to occur,

either by surface runoff from open mill tailings and/or through acid water drainage from

open pits and underground mines (Carvalho et al., 2007). To decrease the radiological

hazard of industrial processes in natural uranium, it is important to manage mining and

processing wastes properly. For instance, inefficient treatment of waste may lead to

uncontrolled transportation of radium. When radium undergoes radioactive decay, it

produces radon gas, which in turn decays to short-lived radionuclides (e.g. 214

Pb and 214

Bi).

These daughters are likely to produce aerosol from tailings located on the surface. Radium

itself tends to associate selectively with minerals, such as barite, even at an extremely low

concentration, and co-precipitation of radium with barium sulphate from acidic solutions is

known (Read et al., 2004).

31

It is possible to reduce radiological contamination from uranium mining (Hore-Lacy and

Cutler, 2004). For example, during the life span of a uranium mine, covering the tailings

with water will minimise the radioactive hazard from physical dispersion. This technique is

mainly applied at mines with higher-grade uranium ores, since lower-grade ores are not

considered a potential hazard at this stage. However, after completion of mining operations,

the tailings will typically be covered by up to 2 m in depth with clay, as a barrier material,

to prevent atmospheric escaping of radon and then a layer of soil to reduce erosion of the

sealing barrier and support the recovery of the plant cover.

1.4 Fractionation of U-series radionuclides in the surface environment

The uranium natural decay series undergoes a series of nuclear transformations starting

from the parent 238

U and ending at stable 206

Pb. During this process, various elements with

different chemical properties are produced, as presented in Fig. 1.5, showing the first part

of U-decay series. Therefore, fractionation in natural waters and sediments occurs as a

result of radionuclide chemistry, weathering processes and interaction between water and

solid, to produce U-series disequilibrium (Chabaux et al., 2008; Jiang et al., 2009). There

are two main mechanisms which lead to fractionation in an aqueous environment. The first

is the difference in chemical properties among U-decay daughters, while the second arises

from physical processes associated with radioactive decay, such as alpha recoil.

32

Figure 1.5 The first part of the 238

U-series

1.4.1 Chemical fractionation

The chemistry of U-Th-Ra in aquatic environments has been reviewed in various

publications (Choppin, 2006; Lehto and Hou, 2010a; Murphy et al., 1999). However,

considerable attention has been focused on the speciation of uranium compared with

thorium or radium; perhaps because it is the most important naturally occurring radioactive

element (Chabaux et al., 2008). Chemical fractionation of U-series products in natural

water is affected by the speciation of radionuclides, the presence of organic matter and

colloids and interaction with solid phases (Chabaux et al., 2008).

33

The first factor affecting the mobility of radionuclides in solutions is the chemical species

present. For instance, solubility of uranium is controlled by its oxidation state. Under

oxidising conditions, uranium forms the soluble U(VI) oxidation state, which is present as

the uranyl species, UO22+

, in aqueous conditions while, in reducing environments,

relatively insoluble U(IV) is dominant. In natural water, the solubility of uranyl ion and its

ability to form diverse complexes with anionic species (e.g. carbonates, phosphates and

many organic functionalities) in water are the main reasons for the number and diversity of

U-minerals (Duff et al., 2002).

By contrast, thorium is one of the least soluble elements. In the U-series, the isotope of

interest is 230

Th, rather than 232

Th, which is the parent of another natural decay series. The

concentration and occurrence of 230

Th depends on its parent 234

U, although, chemically, it is

reported to associate with refractory elements and the resistant fraction in sediments.

Thorium in solution is present in (IV) oxidation state, where it is almost chemically

immobile at low temperatures (Ivanovich and Harmon, 1992).

Radium mass concentrations in natural water are extremely low. For example, 226

Ra

concentration in the Indian rivers were measured (Bhat and Krishnaswamy, 1969) and

found to be in the range of about 23 to 90 ppt (0.8 to 3.0 mBq L-1

), with an average of 45

ppt (1.6 mBq L-1

). However, streams draining limestone, phosphates and U-rich rocks

display higher radium concentration (Porcelli et al., 2001). Radium exists in solution only

in the divalent oxidation state, with relatively limited ability to form complexes, so its

chemical speciation, in natural waters, is simple compared with uranium.

34

The second key influence on chemical fractionation of radionuclides in an aqueous

environment is interaction with organic matter and colloids. Organic matter has an effect on

the uptake of radionuclides onto mineral surfaces. For instance, in a study of uranium

uptake onto pyrite, the presence of dissolved organic matter increased the concentration of

uranium in solution and decreased U(IV) in the solid phase (Bruggeman and Maes, 2010).

The third parameter controlling chemical fractionation of radionuclides in solution is

interaction with solids and mineral surfaces. Removal of actinides from solution to solid

phases is well known, due to reactivity with organic matter and minerals in sediment

surfaces. However, this uptake is element-dependent, as demonstrated for uranium and

thorium (Chabaux et al., 2006; Geibert and Usbeck, 2004). For example, different

mechanisms of uranium uptake onto iron oxide minerals, including adsorption, have been

proposed (Duff et al., 2002). ‘Adsorption’ can be defined simply as a process where one

species is taken up, through physical or chemical processes, on the surface of another.

Adsorption of uranium may include surface complexation, precipitation and incorporation

of uranium in the surface of the host mineral. Both radionuclide concentration and the

presence of microorganisms affect the outcome of these processes (Renshaw et al., 2007).

1.4.2 Physical fractionation

Radioactive decay is the second mechanism leading to fractionation of radionuclides within

the U-series. This process is known as the alpha recoil effect (Fig. 1.6). The recoil effect,

together with the possible displacement of the daughter, contributes to isotopic separation

in water-bedrock interaction. Several studies have reviewed the alpha recoil effect and its

influence on fractionation of 238

U and 234

U in the environment (Bourdon et al., 2003;

Chabaux et al., 2003b; Ivanovich and Harmon, 1992; Skwarzec et al., 2004).

35

As in Figure 1.5, the parent 238

U decays to 234

Th by the emission of an alpha particle. As a

result of momentum conservation, the daughter recoils to a distance in the range of 20-70

nm from the original position; depending on the medium in which the decay takes place.

The medium in which this radioactive decay occurs plays a role in the mobilisation of the

daughter radionuclides. In solid media, such as soils and sediments, the size of the grain

affects recoil fractionation significantly, with a negligible effect in large grain size media

(Chabaux et al., 2003b). The distribution of the parent nuclides in/on the grain and its

surface significantly influence the release of daughter nuclides into solution. Leaching of

the daughter also contributes to fractionation of U-series nuclides (Kronfeld et al., 2004).

Following alpha recoil, the product nuclide will be left in a disturbed crystal structure, due

to local radiation damage. In an aqueous environment, water percolates into microfractures

on the surfaces of mineral grains, enhancing oxidation of 234

U to a higher oxidation state

relative to 238

U. This leads to preferential release of 234

U from damaged lattice sites to

solutions, and hence to disequilibrium (Jiang et al., 2009).

Fig. 1.6 illustrates the processes associated with α-recoil in 238

U (upper right) and the way

in which daughter/ parent activity ratios deviate from, and eventually return to, equilibrium

(the top diagram illustrates trends in 230

Th/234

U versus 234

U/238

U and the bottom one trends

in 230

Th/234

U versus 226

Ra/238

U). Beginning with U-series equilibrium in a grain of bedrock

placed in an aquatic environment due to physical erosion, chemical weathering induces

alteration of the grain composition. As a consequence, and due to the different chemistries

of Ra-Th-U, U-series disequilibrium develops in water and solids (diagonal grey arrows),

and the systems return back to equilibrium by radioactive decay (solid black curves) within

5-7 half-lives of the daughter nuclides. This time required for daughter /parent to return to

36

equilibrium is radionuclide- dependent. In the diagrams of 230

Th-234

U-238

U and 230

Th-226

Ra-

238U, these times are indicated on the decay curve at relevant intervals (in 100 kyr for U-U-

Th system and in 2 kyr for U-Th-Ra system). For example, the system needs about > 1.2

Myr for U-U; about > 400 kyr for U-Th; and about > 8 kyr for Ra-Th to return to

equilibrium due to radioactive decay.

37

Figure 1.6 A diagram summarising alpha recoil effect, and chemical and physical

fractionation of U-series nuclides as a function of time (Dosseto et al., 2008)

38

1.5 U-series in surface waters

During the last three decades, U-series disequilibrium in river components (water,

suspended matter and sediments) has been shown to be a useful tool for various

applications. Many studies of mobilisation of the decay radionuclide during chemical

weathering have been reported (Andersson et al., 1998; Plater et al., 1992; Porcelli et al.,

1997). More recently, the use of U/Th-series isotopes as markers to explore the nature and

time scale of chemical weathering transported by streams in various catchments has been

reviewed (Chabaux et al., 2003a; Dosseto et al., 2006a; Dosseto et al., 2006c; Vigier et al.,

2005). The mobilisation is important, because it can provide essential information relating

to the expected long-term impacts of radioactive waste disposal. Due to the chemical

similarity of the actinides, aquatic systems containing uranium and thorium can serve as

natural laboratories, providing valuable knowledge to support waste disposal.

The uranium concentration in river water was reported to be in a wide range from 0.02

ppm, for the Amazon (Bertine et al., 1970), up to 6.6 ppm, for the Ganges (Bhat and

Krishnaswamy, 1969). About two decades later, a comprehensive study of uranium

concentrations in 40 of the World’s major rivers, as an average, found to be 0.31 ppb

(Palmer and Edmond, 1993). The value decreased to 0.19 ppb when two catchments with

elevated uranium concentrations (the Yellow River and the Ganges-Brahmaputra) were

excluded. A more recent study found uranium concentrations several times higher than

average global levels in the Ganges-Brahmaputra (Chabaux et al., 2001). In contrast, the

average dissolved thorium concentration in river waters is very low (0.1 ppb), while radium

concentrations in rivers are also extremely low as mentioned earlier. In addition to other

factors, differences in solubility between uranium, thorium and radium, contribute to

39

establish U-series disequilibrium. For instance, the 230

Th/234

U activity ratio in river

sediments and solid particles is expected to be greater than unity whereas the same ratio is

likely to be significantly lower than unity in river water (Plater et al., 1992).

1.6 U- series fractionation in river waters

The fate of longer-lived and intermediate half-life radionuclides of the U-series in river

water is influenced by element speciation and processes that affect element transport in an

aquatic system. Chemical weathering of the bedrock, erosion and sedimentation are among

the significant factors affecting the distribution of elements in an aquatic system, and hence

disturbing radioactive equilibrium.

1.6.1 Weathering effect

The concentration of U-Th-Ra in river water depends on lithology of the bedrock

throughout the course of the river. The behaviour of the U-series during chemical

weathering has been identified as the main reason for radionuclide fractionation (Plater et

al., 1988) and, recent studies (Dosseto et al., 2006b; Vigier et al., 2005) have presented the

isotopic ratios of U-series radionuclides in dissolved and suspended components as

evidence to support this assertion. For example, the parent 238

U has a longer half-life

compared with that of the daughter 234

U. In a geological material that is not affected by

chemical weathering (e.g. rock), the ratio 234

U/238

U should be in secular equilibrium.

However, a combination of the alpha recoil effect and chemical weathering leads to a

disequilibrium state. Chemical weathering leaches 234

U from the damaged crystal to the

surrounding water so that 234

U, in an aqueous environment, is expected to be in excess

(Andersen et al., 2009).

40

The isotopic ratio of 234

U/238

U is interpreted as distinguishing between different weathering

behaviours. An excessive physical weathering rate will lead to more 234

U in water due to

the interaction of fresh, small grained material with water, which increases the preferential

leaching of loosely bound 234

U to the water, following alpha recoil. This is consistent with,

for example, the high 234

U/238

U ratio (1.09 - 4.61) found in rivers in the South Island of

New Zealand (Robinson et al., 2004).

In contrast, insignificant disequilibrium of 234

U/238

U from selected rivers from some

Himalayan catchments was claimed, as a result of chemical weathering (Chabaux et al.,

2001). The rationale is that the rapid chemical weathering is associated with the quick and

uniform dissolution of the bulk mineral and is therefore unlikely to cause considerable

fractionation of 234

U from the parent 238

U (Andersen et al., 2009).

1.6.2 Fractionation processes during river transportation

Adsorption of Ra-Th-U by clay minerals, insoluble oxides and the effects of organic matter

on speciation have key roles in the transportation of these radionuclides in natural water.

Organic materials (humic acids and colloids) seem to be more influential than inorganic

sorbents such as Fe-oxides and clay minerals. All these processes affect the content and

behaviour of U-Th-Ra in rivers and will eventually lead to U-series fractionation.

1.6.2.1 Adsorption effects

Uranium and thorium isotopic ratios in the dissolved and particulate compartments of rivers

are assumed to reflect chemical weathering (Plater et al., 1988). As a consequence, residual

fractions are expected to exhibit 234

U/238

U < 1 and 230

Th/238

U > 1. In soluble loads, and

because of leaching, 234

U follows chemical weathering, leading to 234

U/238

U > 1 and

41

230Th/

238U < 1. However, another study by the same author (Plater et al., 1992)

demonstrated that this ideal is often not encountered in reality. In particular, when plotting

234U/

238U versus

230Th/

238U of sediments, isotopic ratios should not appear in a specific

complex area of the diagram, known as the complex zone (Fig. 1.7). The ratios in the

complex zones of the plot should only be found in a complex and weak uranium

accumulation/leaching system over about 25,000 year (Thiel et al., 1983). However, some

U-Th isotopic ratios in sediments from the rivers examined appeared in the complex zone.

The possible reason was adsorption of uranium, with 234

U/238

U > 1, from water onto

sediments. Recently, many studies (Andersson et al., 1998; Chabaux et al., 2003b; Dosseto

et al., 2006c) have demonstrated that minerals resulting from chemical weathering, and

which are common in streams, are effective in adsorbing radionuclides.

Figure 1.7 Edale Valley sediments 234

U/238

U vs 230

Th/238

U diagram as an example of a

complex U-series disequilibrium, with the comlex zone in grey colour

42

1.6.2.2 Speciation and colloidal effects

Organic matter, in addition to clay minerals, is probably the most important sorbent for U-

Th-Ra (Rachkova et al., 2010). Natural organic matter (NOM) results from the degradation

of living organisms. The presence of NOM in rivers is widespread and NOM is an effective

complexant for some U-series radionuclides. The concentration of total organic carbon

(TOC) in organic-rich streams can reach up to 20 ppm, compared with 0.5 ppm in ground

water and sea water. Dissolved, particulate and colloidal organic carbon are all different

components of the total NOM inventory in natural waters. In other words, a large number

of heterogeneous compounds, with various functional groups which are capable of binding

some U-series nuclides, is present in waters (Murphy et al., 1999). Moreover, complex

changes to physicochemical conditions (pH, Eh, ion concentrations) in rivers can also

influence U-Th-Ra fractionation. Several studies have suggested that, in streams with high

organic content, colloidal uranium and its daughters are likely to bond with organic species

(Chabaux et al., 2003b; Porcelli et al., 1997). Conversely, streams draining basaltic areas

often display low organic content. Even in such rivers, there is an effect of organic matter

on uranium transport, although this organic matter appears not to be the main factor

controlling uranium transportation (Pogge von Strandmann et al., 2011; Riotte et al., 2003).

1.7 Natural analogues

Predicting the long-term behaviour of radionuclides in nuclear waste in the environment

surrounding the disposal site is a critical problem facing the nuclear industry. It is possible

to gain useful insights from studying radionuclide transport in systems that are geologically

similar to repositories for radioactive waste disposal (Landa, 2004). Such ‘natural

analogues’ are naturally occurring systems, where processes similar to those expected to

43

occur in nuclear waste facilities are likely to have taken place over a long time. Natural

analogues provide ‘in situ’ laboratories to study how geochemical and biogeochemical

processes affect the migration of contaminants, including the long-lived radionuclides

(Frostick et al., 2011). The data are useful in assessing long-term behaviour and support

both environmental monitoring and radioactive waste management programmes. Moreover,

these results can be extrapolated to transuranium elements, to gain understanding of

nuclear waste repositories (Noseck et al., 2012). Natural uranium deposits have been

studied as analogues for radioactive waste repositories, in particular investigating and

predicting the migration and transportation of contaminants, including radionuclides, to the

surrounding environment (Landa, 2004).

1.7.1 Natural analogues in the UK

In the United Kingdom, many areas have been used as natural laboratories over the last

three decades. The next section highlights briefly three of these sites, describing the origins

of uranium in these sites and providing a summary of the available literature.

1.7.1.1 South Terras U-mine, Cornwall, England

The South-West of the UK, with its scattered granitic intrusions, has a rich history of

mining activities (Gillmore et al., 2001). In particular, Cornwall’s high-temperature (300-

500°C) veins, oriented NE-SW and associated with diverse and complex mineralisation,

have been exploited for different elements, including copper, tin, iron and lead. Several

veins described as low-temperature veins (100- 300°C), cross the high-temperature veins.

These contain a small amount of pitchblende, and have been explored for cobalt, nickel,

iron, lead, uranium and then for radium (Purvis et al., 2004). In the context of radioactive

deposits, the most notable mine in Cornwall is the South Terras mine (50° 20.048ˋ N 4°

44

54.311ˋ W), which was the only UK mine worked primarily for uranium and, subsequently,

radium. The active period of the mine was about 60 years (1870-1930) divided equally

between iron, uranium and radium, although production of tin and iron from the same mine

during the late 19th century has been reported (Smale, 1993). Total production of uranium

from South Terras mine during 1890-1910 was about 739 tonnes (Read et al., 2004).

Uranium was mainly excavated from pitchblende (U3O8) and uraninite (UO2) as primary

ores, but secondary minerals, such as autunite [Ca (UO2)2 (PO4)2.10H2O], zippeite

[(UO2)3(SO4)2(OH)2.8H2O] and torbernite [Cu(UO2)2(PO4)2.8H2O], were reported to be

common in the area (Purvis et al., 2004).

After the cessation of work at the South Terras mine, the spoil heaps and mine buildings,

located beside the River Fal, constituted a natural analogue to allow investigation of the

geochemistry of uranium and its progeny in the vicinity. A previous study (Read and

Hooker, 1992) examined the site as a model for a land-based facility for radioactive waste

disposal. This study focused on leaching and migration of uranium from spoil heaps to

ground water feeding streams flowing toward the River Fal.

1.7.1.2 The Needle’s Eye, Solway, Scotland

The Needle’s Eye is located in the South-West of Scotland on the north coast of the Solway

Firth (National Grid Reference NX 916562). Primary uranium ore, mainly as a pitchblende-

bearing vein, occurs in an ancient cliff at the Southwick coast. The age of this ore was

suggested to be 185 ± 20 Ma (Miller and Taylor, 1966). In specific environments, such as

shallow, organic-rich and reducing environments, the accumulation of uranium is expected

to be quick and effective. This is consistent with behaviour of uranium at Needle’s Eye,

where uranium has accumulated in Quaternary sediments. This rapid accumulation was

45

attributed mainly to three factors (Jamet et al., 1993): uranium is carried by oxidising

ground water from the mineral source, the ancient cliff, and then precipitated. The second

reason is associated with the hydrology of the area. A fault in the vicinity seems to present

a hydraulic barrier to uraniferous groundwater, and on the other side is a flow of ground

water beneath the organic-rich sediments. This sediment provides a reducing environment,

where uranium can precipitate efficiently. The third possible source of uranium

accumulation is leaching of uranium from pitchblende veins located close to the Needle’s

Eye site.

The site has been studied by the British Geological Survey (BGS) since 1985 as a natural

analogue. Extensive studies (mainly documented as technical reports) were conducted to

obtain information about uranium behaviour in the area, in particular, to investigate

retention mechanisms of the uranium in the surrounding organic sediments (Hooker, 1990;

Jamet et al., 1993). The combination of data from solid and ground water indicates that

organic matter in the upper layer of the sediments is the key factor in fixing uranium.

However, in the deeper layers, both organic matter and iron oxyhydroxides contribute to

uranium retention (Hooker, 1990). A few years later, some experimental studies were

conducted to understand the rates of interactions between uranium and Needle’s Eye soils

(Braithwaite et al., 1997; Zhang et al., 1997). The conclusion was that humic substances

with the smallest molecular weight contributed strongly to binding uranium.

1.7.1.3 Edale, Derbyshire, England

Edale is located in Derbyshire, in the E-W trending valley of the river Noe, called the Edale

Valley. The river is fed by small streams flowing from the north and south. Geologically,

the area is situated in an anticlinal structure, the Derbyshire Dome, which exposes a

46

sequence of Carboniferous Dinantian limestones and Namurian deltaic rocks. The bedrocks

of the Edale area are a sequence of fluvio-deltaic mudstones and sandstones of Namurian

age. The sequence varies across the area, reflecting the inhomogeneous nature of the

original sediment supply and channel formation in the Namurian, with mudstones (locally

named shales) representing deeper water facies, and sandstones representing major deltaic

channel sediment deposition. Directly overlying the limestones, and exposed in the valley

floor is the dark mudstone of the Bowland Shale, which is overlain in succession by the

Mam Tor Sandstones, the Shale Grit (coarser sandstone), the Grindslow shales (largely

absent), and the Kinderscout Grits. The Mam Tor Sandstone, Shale Grit and Kinderscout

Grits form the high ground on either side of the Edale Valley. The coarse Shale Grit and

Kinderscout Grits dominate the high ground to the north and west of the Edale Valley, the

latter underlying the plateau of the ‘Dark Peak’ to the north. To the south of the Edale

Valley, the Mam Tor Sandstone, an interbedded sequence of sandstones and shales,

dominates the hill of Mam Tor. Along the Edale Valley, there are also many landslides

have occurred over the late 4000 years, disrupting the succession on the south side of the

Edale Valley. (Dixon and Brook, 2007; Walker, 1966).

During the Quaternary, the softer formations in the sequence have been eroded, and

sediments derived from weathering of the shales, were deposited along the valley. In

particular, the Edale shales are relatively rich in uranium, with concentrations ~50 ppm

(Peacock and Taylor, 1966), and localised higher concentrations associated with phosphatic

nodules at their base (Ford, 1968).

47

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56

Chapter Two

Development of radium separation

The material in the following section has been prepared for submission to the journal

Analyst.

The author designed and performed the radiochemical measurements, analysed the data,

interpreted the results and wrote first draft and the final version of the manuscript.

57

Development of radium separation Siddeeg, S.E., Bryan, N.D., and Livens, F.R.

Centre for Radiochemistry Research, School of Chemistry, University of Manchester,

Manchester, M13 9PL, UK.

2 Introduction

2.1 Radium at South Terras mine

Dispersion of radionuclides from the South Terras uranium mine, Cornwall, to the water

and sediments of the River Fal has been studied (Moliner-Martinez et al., 2004). The

findings suggest that South Terras mine is not a significant source of uranium to the river's

water, despite the nearby location. This is supported by the correlation between the total

cation concentrations and uranium in the surface water of the river, which suggested

background concentrations of uranium. By contrast, the concentration of uranium in

sediment immediately beneath the outflow pipe was extremely high (up to 1000 ppm).

Another study has been conducted focusing on two abandoned metalliferous mines in

Cornwall and Devon (Gillmore et al., 2001). This study demonstrated that the

concentrations of radon gas can be extremely high (exceeding the permissible limit),

particularly at South Terras. The maximum dose rate at South Terras is 18 mSv h-1

, which

is extremely high compared with the relevant annual limit (1 mSv yr-1

). The measurements

were performed from 1992-1994 and again in 2000, at different locations around the mine.

The results revealed no significant difference regarding the location of the measurement

(e.g. in the main entrance or at 2 metres outside the mine). However, the results showed

radon concentrations are seasonally dependent, being slightly higher in winter compared

with spring. Measurements of radium in the vicinity of South Terras were therefore likely

58

to be of interest, so a method for radiochemical separation and alpha spectrometry was

devised, as described below.

2.2 Techniques for radium measurement

Radium has about 20 radioactive isotopes, of which the most well-known are the four

natural radium isotopes: 223

Ra, 224

Ra, 226

Ra and 228

Ra are summarised in Table 2.1

(Gillmore et al., 2001; Jia and Jia, 2012). This allows diverse analytical techniques to be

employed for radium measurement in environmental samples.


radium isotopes

Isotope Decay series Decay mode Half-life Energy (MeV)

223Ra Actinium alpha 11 d 5.72 (51.6%) 5.61 (25.2 %)

224Ra Thorium alpha 3.6 d 5.69 (94.9 %) 5.45 (5.1 %)

226Ra Uranium alpha 1.6 x 10

3 y 4.78 (94.5 %) 4.60 (5.6 %)

228Ra Thorium beta 5.8 y 0.046 Emax

The selection of the analytical technique to be used depends on several factors. These

include parameters such as sample matrix, the presence of accompanying radionuclides, the

detection limit required and, from an economic perspective, availability and capital cost.

The most common conventional radiometric methods with which to measure radium

isotopes in the environmental samples are alpha spectrometry, gamma spectrometry, liquid

scintillation counting and radon emanation (Benedik et al., 2010). Atom counting methods,

specifically inductively-coupled plasma mass spectroscopy (ICP-MS), thermal ionisation

59

mass spectroscopy (TIMS) and accelerator mass spectroscopy (AMS) methods can also be

used (Hou and Roos, 2008).

2.2.1 Radiometric techniques

Radiometric methods are the common techniques for radium determination in the

environmental samples. Some techniques, such as alpha spectrometry, gamma spectrometry

and liquid scintillation are based on direct measurement of Ra decays. Others, for example

gamma spectrometry or liquid scintillation of radium daughters, or the radon emanation

technique, are indirect methods. The following sections will describe briefly the three most

routine radiometric methods.

2.2.1.1 Alpha spectrometry

Alpha spectrometry, using solid state semiconductor detectors, provides a direct method for

radium isotope measurement. The approach, briefly, involves three stages: sample

treatment, chemical separation and source preparation. In principle, alpha spectrometry

offers various advantages over alternative radiometric techniques. Firstly, chemical

separation purifies the analyte from the matrices, resulting in higher sensitivity. Secondly,

if one allows ingrowth of 228

Ra decay products, alpha spectrometry provides the possibility

of measuring simultaneously the four naturally-occurring radium isotopes in the same

prepared source. In particular for 228

Ra, the prepared alpha disk should be stored for 6

months to allow the ingrowth of 228

Th, which can be measured by analysis of the 224

Ra

peak at 5.69 MeV. Thirdly, by using a radiotracer, the efficiency of the chemical separation

can be evaluated (internal calibration). Additional advantages are the low detection limits

(0.2-0.5 mBq at 2 days counting time) of the detectors and the low background (Jia and Jia,

2012). Therefore, alpha spectrometry is one of the preferred techniques to measure radium

60

isotopes in environmental samples (Jia and Jia, 2012; Zhang et al., 2009). However, the

drawbacks of alpha spectrometry include: a requirement for proper radiochemical

separation of the analyte from the matrix, availability of a radiotracer to be used for yield

determination, and good source preparation before introducing the planchette to the

detector for counting. The separation demands tedious chemistry, and a suitable isotope of

the analyte, or a chemical analogue, can be used as the tracer. For alpha source preparation,

different methods (e.g. co-precipitation and electrodeposition) can be used. All these factors

make alpha spectrometry time consuming (Smith and Mercer, 1970).

The availability of radiochemical separation methods and radiotracers (e.g. 225

Ra and 113

Ba)

for 226

Ra has made alpha spectrometry the preferred technique (Blanco et al., 2002; Vasile

et al., 2010). However, because 133

Ba is an analogue of radium, 225

Ra is often the best

option for use as a tracer, although it does require complex calculations. 225

Ra (t1/2 14.8 d) is

a β-emitter, decaying to 225

Ac (t1/2 10.0 d), which in turn decays to the short-lived α-

emitters217

At and 213

Po. The calculation of the chemical recovery is based on the counts of

217At at 7.07 Mev or of

213Po at 8.38 Mev. The prepared alpha source is counted twice,

firstly just after electrodeposition to get the maximum counts of 226

Ra, and again after 17

days when the activity of 225

Ac reaches a maximum.

The radium spectra obtained from an alpha source prepared in this way are complicated by

the presence of various peaks from 226

Ra and 225

Ra progeny 226

Ra itself produces two

peaks at 4.60 and 4.78 MeV in addition to peaks of 222

Rn at 5.60 MeV, 218

Po at 6.87 MeV

and 214

Po at 7.88 MeV. The peaks from 225

Ra decay include those of 225

Ac at 5.94 MeV,

221Fr at 6.30 Mev,

217At at 7.07 MeV and

213Po at 8.38 MeV. However, semiconductor

61

detectors with counting efficiency 20- 35% provide the resolution and sensitivity to

interpret these complex spectra (Blanco et al., 2002; Crespo and Jimenez, 1997; Sill, 1987).

2.2.1.2 Gamma spectrometry

When low detection limits are not necessary, as in soil and sediment samples with elevated

concentrations of radium, gamma ray spectrometry is an attractive technique. The

technique uses semi-conductor detectors (high purity germanium detector, HPGe), cooled

down to liquid nitrogen temperature (~ 77 K). It is non-destructive, does not require

chemical separation, and multi-nuclide determinations can be made with a suitable amount

of sample. These features are considered advantageous for gamma spectrometry over alpha

spectrometry (Zhang et al., 2009). However, the relatively poor limit of detection (0.1-1.0

Bq at 5 hours counting time), and the need for energy calibration, efficiency calibration and

careful matching of matrix and of sample geometry all limit the usefulness of gamma

spectrometry.

Direct measurement of 226

Ra can be achieved using a gamma line at 186.2 keV. However,

the gamma ray abundance for this line is only 3.6%, and the 235

U line at 185.7 keV will

interfere. Therefore, to obtain accurate radium concentrations from direct gamma

spectrometry, it is necessary to consider correction for the added counts from 235

U, or to

separate radium from uranium before counting (Vasile et al., 2010). In contrast, indirect

measurement of 226

Ra by gamma spectrometry can be performed after ensuring secular

equilibrium with the short-lived progeny (214

Pb and 214

Bi) of its direct daughter 222

Rn. The

method is accurate, but it requires at least four weeks waiting time for 222

Rn build-up to

reach secular equilibrium with its daughters 214

Pb and 214

Bi, in a well-sealed plastic

container. The activity concentration of 226

Ra can then be calculated from the photopeaks at

62

295 keV and 351 keV for 214

Pb and 609, 1120 and 1764 keV for 214

Bi (Kohler et al., 2002;

Vasile et al., 2010).

2.2.1.3 Liquid scintillation counting

Liquid scintillation is a counting method based on using photo multiplier detectors, and can

be either qualitative or quantitative. It is an alternative, simple, fast and sensitive method

(detection limit 0.3-1.4 mBq at 6 hours counting time) for measuring 226

Ra in a large

number of environmental samples. As in gamma spectrometry, determination of 226

Ra by

liquid scintillation can be either direct, or indirect, using radium progeny. The more

common approaches, though, are based on the solubility of radon gas in an organic

scintillation cocktail, followed by measuring alpha emissions from the daughter 214

Po, or

the sum of 226

Ra, 222

Rn, 218

Po and 214

Po. However, it is essential to allow sufficient time to

establish secular equilibrium with radon and its short-lived progeny by keeping the sample

in a closed vial for around 3-4 weeks before counting (Al-Masri and Blackburn, 1996). For

natural water samples, the radium concentration is expected to be very low so, to improve

measurements, pre-concentration might be employed. For instance, radium co-precipitation

with lead/barium sulphate followed by dissolution in alkaline EDTA, or coprecipitation on

manganese dioxide has been reported (Cooper et al., 1988; Nour et al., 2004).

2.2.2 Atom Counting Techniques

Recently, ICP-MS, TIMS and AMS techniques have been employed to measure radium in

different environmental samples (Hou and Roos, 2008; Lariviere et al., 2006; Pietruszka et

al., 2002; Sharabi et al., 2010). Compared with radiometric methods, atom counting

techniques have many potential advantages; notably, a small volume of the sample is

required, the analysis can be more precise and accurate, and the analysis time can be

63

shorter. However, isobaric interference and blank correction of most of them need to be

considered (Lariviere et al., 2006).

Among the mass spectrometric techniques, ICP-MS is most commonly applied in 226

Ra

measurements. The method for an aqueous sample is direct, while dissolution may be

employed for solid samples. The detection limit varies depending on the sample matrix, but

is generally in the range of 0.22 – 11 Bq/L (Epov et al., 2003). The drawbacks of ICP-MS

include matrix effects, formation of isobaric species, mass fractionation during detection,

formation of multiple ions, polyatomic atomic interference and the abundance sensitivity

with respect to neighbouring matrix ions (Sharabi et al., 2010). In TIMS, the similar

ionisation potentials of Ba (5.21 eV) and Ra (5.28 eV), make radium difficult to analyse

properly (Jia and Jia, 2012). One approach to reduce problems in mass spectrometric

methods is to separate radium in the environmental samples from the matrix. Although this

can decrease the effort in mass spectrometric analysis of radium, it requires specialist

facilities and methods. These restrictions, combined with the high cost of mass

spectrometric instruments mean that such measurements are less common .

2.3 Methodology

2.3.1 Preparation of radium test solution

The original 226

Ra solution was supplied by AEAT Harwell and diluted as required. For this

work, a radium working solution of 2 Bq mL-1

was prepared in a 25 ml flask from a 226

Ra

solution (Ra226KD2) with a total activity of 11.75 kBq. Starting from the Ra226KD2

solution (1 kBq mL-1

), 50 l was transferred by micropipette into a 25 ml volumetric flask,

and made up to volume with 2 M nitric acid to give a total activity of 50 Bq. This solution

was designated Ra226SS1.

64

2.3.2 Radiochemical separation testing

The radiochemical separation in this study follows the procedure developed by Aguado et.

al. (2008). The method given here is based on a series of tests carried out to optimise the

separations, which are described below. Fifty mL 0.1 M HNO3 was placed in a 150 mL

beaker and 0.25 mL from solution Ra226SS1 was added and mixed thoroughly. 1 mL of

conc. H2SO4 was added and mixed, followed by 2 g solid K2SO4, which was allowed to

dissolve. Then, 1 mL of 0.24 M Pb(NO3)2 was added drop-wise, to form lead sulphate. The

suspension was left to precipitate for 10 minutes, and then separated by centrifugation in a

50 mL tube at 3000 rpm (about 6200 g) for 10 minutes). The precipitate was washed with

20 mL 0.2 M H2SO4/ 0.1 M K2SO4.

The Pb(Ba,Ra)SO4 precipitate was dissolved completely with gentle heating by adding 4-5

mL of 0.1 M ethylenediaminetetraacetic acid (free acid form; H4EDTA), previously

adjusted to pH 10 using concentrated ammonia solution. The solution was passed through

an anion exchange column (Bio-Rad AG1-X8, 100-200 mesh, chloride form, 5 cm x 6 mm)

to remove sulphate, and the column was then washed with 13 mL EDTA/ammonia

solution. Thorium, sulphate and actinium are retained on the column. To the eluted

solution, 1 mL 5 M CH3CO2NH4 was added and the pH adjusted with 8 M HNO3 to 4.5.

A cation exchange column (Bio-Rad AG50W-X12, 200- 400 mesh, 8 cm x 7 mm) was

converted to the ammonium form by passage of 15 mL 1.5 M CH3CO2NH4, then

conditioned with 15 mL of 0.25 M CH3CO2NH4, previously adjusted to pH 4.5 using 8 M

HNO3. The radium and barium fraction was transferred to the column at a low flow

(~1mL/minute). The residual Th, Pb and Ac were eluted by washing the column with 50

mL 1.5 M CH3CO2NH4 in 0.1 M HNO3. Ba was eluted by washing the column with 40 mL

65

2.5 M HCl. Finally, Ra was eluted with 25 ml 6 M HNO3 and the effluent dried using a heat

lamp.

2.3.3 Alpha source preparation

For electrodeposition, a cell was prepared and checked for leakage. The Ra fraction was

dissolved in 0.5 mL 0.1 M HNO3, and the solution transferred to an electrodeposition cell

with 1 mL 0.05 M HCl and 9 mL ethanol. To perform electroplating, the cell current was

maintained at 120 mA and the voltage at 25 V during the 30-minute process. Two drops of

concentrated ammonia were added one minute before the end of the process. The cell was

dismantled and the planchette removed with tweezers. The planchette was dried at room

temperature and a sealed plastic bag was used to carry the prepared alpha planchette to the

counting room.

2.3.4 Sample measurement

The samples were counted using multiple Canberra Model 7401 alpha spectrometers

connected with a Canberra Model 2000 Integrated ADC-Mixer/Router, counting into 512

channels. Each chamber was equipped with a PIPS detector, with 450 mm2 active area and

100 µm depletion depth. The system was connected to a personal computer, and a Canberra

system Multichannel Analyser was employed to analyse the spectrum. Counting time

varied for each sample, but in order to obtain adequate counting statistics, was set up to 20

hours for each sample. Radium-226 was quantified using its two distinct alpha peaks at

4.784 MeV (94.04 %) and 4.601 MeV (5.95%).

66

The procedure of radium separation and alpha source preparation for spike solutions was

tested, and minor modifications to the method were made to optimise conditions for radium

chemical recovery. These are presented in the results and discussion section.

2.4 Results and discussion

The calculation was performed in the same way for all counted alpha sources of 226

Ra,

based on the activity of 226

Ra used during the radiochemical separation and an assumed

efficiency of 25% for the PIPS detector.

The formula used to calculate the recovery of 226

Ra in the prepared alpha sources is:

( ) ( )

Table 2.2 summarises the modifications made to the method, the purpose of the change, the

net counts from the alpha spectrum, the counting time and the chemical recovery for the

samples.

67

Table 2.2 Optimising radium separation conditions in spiked nitric acid test samples and 226

Ra chemical recovery

Sample ID What was done Purpose counts Counting time (s) Yield %

Radium T1 The protocol was tested, but the

electrolysis conditions were not

applied properly as a fault of the

power supply

To examine the separation

and the electrodeposition

353 ± 18 83274 3.4 ± 5.3

Radium T2 PbSO4 precipitate was washed

with 0.2 M H2SO4/0.1 M K2SO4.

Electroplating conditions were

applied properly

To test the effect of

removing the extra H2SO4

in the coprecipitation step

1630 ± 40 16000 81.5 ± 2.5

Radium T3 Reducing the amount of

chemicals to half

To be consistent with the

added 0.1 M HNO3

1060 ± 32 16000 53 ± 3.1

Radium T4 No more changes to the protocol

other than increasing the counting

time

To improve the counting

statistics

4490 ± 67 72000 49.9 ± 1.5

Radium T5 The amounts of H2SO4 and K2SO4

were doubled and the counting

time increased

To test the effect of SO42-

on PbSO4 precipitate

4041 ± 63 72000 44.9 ± 1.6

Radium T6 The separation of PbSO4 was

performed at about 1050C, and

0.05 M HCl was not added to the

electrolyte solution

To test the separation of

PbSO4 by heating, and test

the electrodeposition

3710 ± 60 72000 41.2 ± 1.6

68

Radium T7 Washing the PbSO4 precipitate

with 0.2 M H2SO4/0.1 M K2SO4

was skipped and heating was used

for its separation

To test the efficiency of

the suspended separation

and the effect of H2SO4

7580 ± 87 72000 84.2 ± 1.1

Radium T8 The amounts of 0.1 M nitric acid,

sulphuric acid and lead nitrate

were doubled with no wash to

PbSO4

To test the efficiency of

the suspended separation

and the effect of H2SO4

7310 ± 85 72000 81.2 ± 1.2

Radium T9 The PbSO4 precipitate was

washed with 0.2 M H2SO4/0.1 M

K2SO4

To test the effect of

purifying the precipitate

5860 ± 75 72000 65.1 ± 1.3

Radium 10 Same as Radium T9 To test the effect of

purifying the precipitate

6334 ± 79 72000 70.4 ± 1.3

Radium T11 The wash of the PbSO4 with 0.2

M H2SO4/0.1 M K2SO4 was

skipped. No use of 0.05 M HCl in

the electrolyte

To test the effect of H2SO4

on purifying PbSO4, and to

test electrolyte solution

3550 ± 59 72000 39.4 ± 1.7

Radium T12 The amounts of the sulphuric acid

and the lead nitrate were doubled

To test all modifications

done so far

7763 ± 88 72000 86.3 ± 1.1

69

From Table 2.2, other than the low yield from first experiment, which was linked to a

fault in the power supply during electroplating, the chemical recovery of the radium was

in the range of 40 – 86 %, with a mean value of 63%. This yield is in a good agreement

with the literature (Aguado et al., 2008; Blanco et al., 2002; Crespo and Jimenez, 1997),

and the method is relatively insensitive to the changes. The modifications applied to the

method, such as the order of adding the reagents to precipitate the lead sulphate, the step

of washing the precipitate and the preparation of the electrolyte solution for

electrodeposition, seem to have no major effect on the radium recovery. However, the

two lowest yields obtained, T6 and T11, were associated with the absence of weak

hydrochloric acid in the electrolyte solution. The presence of excess hydrogen ion

during the electroplating step has been recommended, since radium is electroplated

from organic solution (Aguado et al., 2008)

70

2.5 Conclusions and recommendations

From the experiments performed to optimise radium separation, the results suggested

the following could be the best conditions to obtain higher recovery of the radium:

Using 2 mL of 0.05 M HCl in the electrolyte solution improves deposition

of radium onto the planchette during electrodeposition. This amount makes

the electroplating process proceed in a constant current of 0.12 A at 30 V.

Washing the PbSO4 precipitate with a mixture of 0.2 M H2SO4 and 0.1 M

K2SO4 before dissolution is important to obtain a clean precipitate before

proceeding to the next steps.

Finally, count for a long time (> 20 hours) to obtain enough net counts in

order to reduce the counting error.

71

References

Aguado, J.L., Bolivar, J.P., García-Tenorio, R., 2008. 226

Ra and 228

Ra determination in

environmental samples by alpha-particle spectrometry. Journal of Radioanalytical and

Nuclear Chemistry 278, 191-199.

Al-Masri, M.S., Blackburn, R., 1996. Radioanalytical methods for determination of

alpha emitters in the environment. Radiation Physics and Chemistry 47, 171-175.

Benedik, L., Repinc, U., Strok, M., 2010. Evaluation of procedures for determination of

Ra-226 in water by α-particle spectrometry with emphasis on the recovery. Applied

Radiation and Isotopes 68, 1221-1225.

Blanco, P., Lozano, J.C., Tome', F.V., 2002. On the use of 225

Ra as yield tracer and

Ba(Ra)SO4 microprecipitation in 226

Ra determination by α-spectrometry. Applied

Radiation and Isotopes 57, 785-790.

Cooper, E.L., Brown, R.M., Milton, G.M., 1988. Determination of 222

Rn and 226

Ra in

environmental waters by liquid scintillation counting. Environment International 14,

335-340.

Crespo, M.T., Jimenez, A.S., 1997. On the determination of radium by alpha-

spectrometry. Journal of Radioanalytical and Nuclear Chemistry 221, 149-152.

Epov, V.N., Lariviere, D., Evans, R.D., Li, C., Cornett, R.J., 2003. Direct determination

of 226

Ra in environmental matrices using collision cell inductively coupled plasma

mass-spectrometry. Journal of Radioanalytical and Nuclear Chemistry 256, 53-60.

Gillmore, G.K., Phillips, P.S., Pearce, G., Denman, A., 2001. Two abandoned

metalliferous mines in Devon and Cornwall, UK: radon hazards and ecology.

International radon symposium, 94-105.

72

Hou, X., Roos, P., 2008. Critical comparison of radiometric and mass spectrometric

methods for the determination of radionuclides in environmental, biological and nuclear

waste samples. Analytica Chimica Acta 608, 105-139.

Jia, G., Jia, J., 2012. Determination of radium isotopes in environmental samples by

gamma spectrometry, liquid scintillation counting and alpha spectrometry: A review of

analytical methodology. Journal of Environmental Radioactivity 106, 98-119.

Kohler, M., Preube, W., Gleisberg, B., Schafer, I., Heinrich, T., Knobus, B., 2002.

Comparison of methods for the analysis of 226

Ra in water samples. Applied Radiation

and Isotopes 56, 387-392.

Lariviere, D., Taylor, V.F., Evans, R.D., Cornett, R.J., 2006. Radionuclide

determination in environmental samples by inductively coupled plasma mass

spectrometry. Spectrochimica Acta Part B: Atomic Spectroscopy 61, 877-904.

Moliner-Martinez, Y., Campins-Falco, P., Worsfold, P.J., Keith-Roach, M.J., 2004. The

impact of a disused mine on uranium transport in the River Fal, South West England.

Journal of Environmental Monitoring 6, 907-913.

Nour, S., El-Sharkawy, A., Burnett, W.C., Horwitz, E.P., 2004. Radium-228

determination of natural waters via concentration on manganese dioxide and separation

using Diphonix ion exchange resin. Applied Radiation and Isotopes 61, 1173-1178.

Pietruszka, A.J., Carlson, R.W., Hauri, E.H., 2002. Precise and accurate measurement

of 226

Ra-230

Th-238

U disequilibria in volcanic rocks using plasma ionization

multicollector mass spectrometry. Chemical Geology 188, 171-191.

Sharabi, G., Lazar, B., Kolodny, Y., Teplyakov, N., Halicz, L., 2010. High precision

determination of 228

Ra and 228

Ra/226

Ra isotope ratio in natural waters by MC-ICPMS.

International Journal of Mass Spectrometry 294, 112-115.

73

Sill, C.W., 1987. Determination of radium-226 in ores, nuclear wastes and

environmental samples by high-resolution alpha spectrometry. Nuclear and Chemical

Waste Management 7, 239-256.

Smith, K.A., Mercer, E.R., 1970. The determination of radium-226 and radium-228 in

soils and plants, using radium-225 as a yield tracer. Journal of Radioanalytical

Chemistry 5, 303-312.

Vasile, M., Benedik, L., Altzitzoglou, T., Spasova, Y., Watjen, U., Gonza'lez de

Orduna, R., Hult, M., Beyermann, M., Mihalcea, I., 2010. 226

Ra and 228

Ra determination

in mineral waters-Comparison of methods. Applied Radiation and Isotopes 68, 1236-

1239.

Zhang, W., Ungar, K., Chen, J., St-Amant, N., Tracy, B.L., 2009. An accurate method

for the determination of 226

Ra activity concentrations in soil. Journal of Radioanalytical

and Nuclear Chemistry 280, 561-567.

74

Chapter Three

Dispersion of U-series radionuclides in stream sediments from

Edale Valley, UK

The material in the following section was given as an oral presentation at The Co-

ordinating Group for Environmental Radioactivity (COGER) meeting, 02- 04 April

2012, in Portsmouth. This manuscript has been prepared for submission to Journal of

Environmental Science: Processes & Impacts.

The author was involved in the collection of samples, performed the radiochemical

measurements, analysed the data, interpreted the results and wrote first draft and the

final version of the manuscript.

75

Dispersion of U-series radionuclides in stream sediments from

Edale Valley, UK

Siddeeg, S.E., Bryan, N.D., and Livens, F.R.



Abstract

The spatial distribution of 238

U-series radionuclides, specifically 238

U, 234

U, 230

Th and

226Ra, has been determined in stream sediments from the Edale valley, Derbyshire,

United Kingdom, to explore the behaviour of U-series radionuclides during weathering.

For uranium and thorium, two different extraction methods were used, total dissolution

with HNO3/HF in a microwave and leaching with aqua regia. This was followed by

radiochemical separation using extraction chromatography, then alpha spectrometry

measurement. The total radium contents in the sediments were measured using gamma

spectrometry, while the leached fraction was measured in the same way as for uranium

and thorium. The total sediment radionuclide content of uranium and thorium ranges

from ~10 up to ~200 Bq.kg-1

, while the radium activity concentration is highly variable,

from ~15 up to 180 Bq.kg-1

. In the aqua regia extractions, the uranium and thorium

contents are in the range of ~5 to ~100 Bq.kg-1

, while the radium activity concentrations

are similar to those measured by total dissolution. All the radionuclides show no

correlation with organic matter content. The activity ratios 234

U/238

U, 230

Th/238

U and

226Ra/

238U were used to determine the degree of radioactive equilibrium. The data show

disequilibrium in most of the sediments, with activity ratios of 234

U/238

U, 230

Th/238

U and

226Ra/

238U > 1, inconsistent with evolution through straightforward weathering

processes. Multivariate cluster analysis based on five variables, the activity

concentrations of 238

U, 234

U, 230

Th, 226

Ra and loss on ignition, was employed to group

76

the data and identify five distinct clusters. There seems to be a link between high

radionuclide concentrations and proximity to landslips.

3 Introduction

3.1 Naturally occurring uranium

Uranium is a radioactive element with three naturally occurring isotopes (234

U, 235

U and

238U). Among the three radioisotopes,

238U (t1/2= 4.468 x 10

9 y ;> 99.2 atom %) is the

parent of one of the three natural decay series. The daughter radionuclides with

intermediate half-lives 234

U (t1/2= 2.48 x 105 y),

230Th (t1/2= 7.52 x 10

4 y) and

226Ra (t1/2=

1.6 x 103 y) are crucial in studying U-series disequilibria (Ivanovich and Harmon,

1992). As uranium both represents a significant component of radioactive wastes and

may serve as an analogue for other actinides, its behaviour is interesting from a waste

disposal point of view (Pekala et al., 2010)

Uranium may exist in several oxidation states, with tetravalent and hexavalent being the

most dominant in the environment. In oxic systems, such as river water and surface

sediments, the higher oxidation state is favoured, whereas in anoxic environments the

lower oxidation state is common (Michel, 1984). In the hexavalent oxidation state,

uranium is relatively more soluble and mobile depending on pH and redox conditions.

This mobility may result either from complexation with different ligands (e.g.

carbonates and hydroxides) or from binding to colloids in organic-rich waters (Chabaux

et al., 2003; Plater et al., 1992).

Thorium in the environment, including radiogenic 230

Th, predominantly exists in the

tetravalent oxidation state. It is insoluble and immobile in aqueous media at low

temperature and pH > 3 (Plater et al., 1992). However, in the presence of organic matter

(e.g. humic and fulvic acids) and minerals, such as hematite, thorium may be mobilised

due to complexation (Murphy et al., 1999). Consequently, it is expected that, in waters,

77

the 230

Th/234

U ratio will be significantly lower than unity, whereas in sediment this ratio

is expected to be greater than unity (Dosseto et al., 2008).

Radium belongs to the alkaline earth metals and exhibits only the divalent oxidation

state. Compared to uranium and thorium, it is highly reactive and can easily adsorb onto

mineral surfaces and/or replace calcium in minerals. The radium distribution in

weathering profiles is less well-documented than that of thorium and uranium.

However, the cycling of radium by plants and its retention by organic matter in the soil

profile may contribute to enrichment of Ra relative to Th and U (Chabaux et al., 2003).

Erosion and chemical weathering modify rocks, and rivers enhance migration by

carrying uranium away as part of the soluble phase, suspended matter or as sediments

(Chabaux et al., 2008). Rock mineralogy and water-solid interactions both lead to

redistribution and transport of elements leached from rocks. In particular, colloids,

organic matter and different mineral phases have a significant effect on uranium

transport in the surficial environment (Pogge von Strandmann et al., 2011; Suresh et al.,

2011). Therefore, studying natural radionuclides in stream sediments provides insight

into their sources, behaviour and fate along the river course.

3.2 Fractionation of 238

U-series

In a geological system, which has been closed for a sufficiently long time (ca 1.5 Ma),

238U-series isotopes tend to be in secular equilibrium so that the activity concentrations

of the parent (238

U) and the intermediate-lived daughters (234

U, 230

Th and 226

Ra) are

essentially equal. However, in the surface environment and because of the varied half-

lives of the daughters, differences in the daughters' chemistry and presence of organic

and inorganic colloids, radioactive disequilibrium is likely to be dominant (Andersson

et al., 1998; Dosseto et al., 2008; Plater et al., 1992). Chemical weathering and water-

rock interactions enhance this fractionation, and once it takes place, it may last for

millions of years (Chabaux et al., 2008; Noseck et al., 2012).

78

In the case of the uranium isotopes (238

U and 234

U), 234

U tends to exhibit a greater

concentration in the soluble phase (Dosseto et al., 2008) due to:

i) ejection of 234

Th from the grain to the surrounding liquid when its recoil range is

greater than the distance to the grain boundary, followed by the decay of the short half-

life 234

Th (t1/2 = 24 days) to 234

U;

ii) damage of the crystal lattice of the mineral grain from -decay events, which

enhances the escape of the product nuclide from the damaged site (Dosseto et al., 2008).

The result is an expected 234

U/238

U activity ratio greater than unity in waters and less

than unity in river sediments (Andersen et al., 2009; Vigier et al., 2006).

Additional fractionation, associated with chemical properties of the radionuclide in

natural water, is expected in the U-series. The fractionation between Ra-Th-U occurs

since thorium is insoluble and tends to become associated with the solid phase, while

uranium is more soluble in most surface environments and radium relatively displays

intermediate solubility (Chabaux et al., 2003). Thus, for example, the 230

Th activity

concentration in sediments will depend both on the chemistry of its direct precursor

(234

U) and interaction with the surrounding environment (Galindo et al., 2007). The

presence of colloids, particularly organic, also influences U-series fractionation through

complexation of Ra, Th and U that affects their mobility and modifies their distribution

(Andersson et al., 1998).

3.3 Objectives of the study

This study investigates the spatial distribution of 238

U-series radionuclides in sediments

from Edale, focusing on 238

U, 234

U, 230

Th and 226

Ra, to understand the behaviour and

mobility of U-series radionuclides during weathering by identifying parameters

controlling this mobility. This work aims to explore the phases of the minerals and/or

organic matter involved in the retention of the U-series radionuclides and the level of

association between these variables.

79

3.4 Materials and methods

3.4.1 The study area

Edale is located in Derbyshire, in the E-W trending valley of the river Noe, called the

Edale Valley. The river is fed by small streams flowing from the north and south.

Geologically, the area is situated in an anticlinal structure, the Derbyshire Dome, which

exposes a sequence of Carboniferous Dinantian limestones and Namurian deltaic rocks.

The bedrocks of the Edale area are a sequence of fluvio-deltaic mudstones and

sandstones of Namurian age. The sequence varies across the area, reflecting the

inhomogeneous nature of the original sediment supply and channel formation in the

Namurian, with mudstones (locally named shales) representing deeper water facies, and

sandstones representing major deltaic channel sediment deposition. Directly overlying

the limestones, and exposed in the valley floor is the dark mudstone of the Bowland

Shale, which is overlain in succession by the Mam Tor Sandstones, the Shale Grit

(coarser sandstone), the Grindslow shales (largely absent), and the Kinderscout Grits.

The Mam Tor Sandstone, Shale Grit and Kinderscout Grits form the high ground on

either side of the Edale Valley. The coarse Shale Grit and Kinderscout Grits dominate

the high ground to the north and west of the Edale Valley, the latter underlying the

plateau of the ‘Dark Peak’ to the north. To the south of the Edale Valley, the Mam Tor

Sandstone, an interbedded sequence of sandstones and shales, dominates the hill of

Mam Tor. Along the Edale Valley, there are also many landslides have occurred over

the late 4000 years, disrupting the succession on the south side of the Edale Valley.

(Dixon and Brook, 2007; Walker, 1966).

During the Quaternary, the softer formations in the sequence have been eroded, and

sediments derived from weathering of the shales, were deposited along the valley. In

particular, the Edale shales are relatively rich in uranium, with concentrations ~50 ppm

(Peacock and Taylor, 1966), and localised higher concentrations associated with

phosphatic nodules at their base (Ford, 1968).

80

3.4.2 Sampling and sample pretreatment

A total of 25 samples of stream sediment were collected from the Vale of Edale (Fig.

3.1), during two field trips (10 and 17 December 2010). They were saved in paper Kraft

envelopes for further analysis in the laboratory. The samples were wet sieved through a

2 mm sieve, before being left to air dry in the laboratory. Once dry, they were

disaggregated using a mortar and pestle and homogenised before chemical treatment.

Approximately 1- 2 g of each sample, accurately weighed, was heated to 550 °C for 5

hours and loss on ignition measured in three replicate.

3.4.3 Mineralogy of the samples

The mineralogy of the sediments was examined qualitatively by XRD using an X’Pert

Powder (Cu Kα 0.152 nm, 40 kV, 30 mA) diffractometer equipped with a Multi-

Channel Detector (X’Celerator). The samples were sieved (80 mesh) and appropriate

amounts (~0.5 g) were placed onto the sample holder. A smooth flat surface was

obtained before introducing the specimen to the instrument. The exposure time was 30

minutes, and the phase identification for the collected spectrum was performed using

the X’Pert HighScore Plus powder pattern analysis tool.

3.4.4 Radiochemical characterisation

3.4.4.1 Sediment dissolution

For total dissolution of the sediments, 0.2 g of the sediment was ashed in a muffle

furnace at 550⁰ C for 5 hours, and placed in a closed vessel and wetted overnight with a

mixture of 1.0 mL deionised water, 3.0 mL concentrated nitric acid and 6.0 mL

concentrated hydrofluoric acid. The sample was then digested in a microwave oven with

ramping time 10 minutes to 140 ⁰C (~150 psi) and 50 minutes holding time, and this

was repeated three times, before evaporation. Finally, 2.0 mL of 20 % nitric acid was

added to the residue and the volume was made up to 20 mL with deionised water.

81

For leaching, a known amount of the sediment (from 0.5 – 2.0 g) was ashed and then

leached with 15.0 mL aqua regia (concentrated hydrochloric and nitric acids in a 3:1

ratio) at near boiling point for 3 hours. The aim was to extract the radionuclides which

were not associated with primary minerals. In particular, the leached fractions include

those associated with organic matter and adsorbed onto the surfaces of minerals and

secondary phases (Marsden et al., 2001; Pekala et al., 2010). After the leaching was

evaporated, the volume was made up to 50 mL using 0.1 M HNO3.

3.4.4.2 Uranium and thorium separation

The experimental setup was based on extraction chromatography and modified from

that proposed to separate Th/U in geological samples (Carter et al., 1999). To the ashed

sediment, spikes of 232

U and 229

Th (40 mBq and 50 mBq respectively), prepared from

certified standard solutions (CERCA LEA, France and National Physical Laboratory,

U.K.) were added. To the sample solution, whether produced by total dissolution or

leaching, 1 x 5 mL portion of concentrated HNO3 was added and the solution taken to

near dryness under a heat lamp. The nitric acid treatment was repeated. The residue was

dissolved in 10 mL of 3 M HNO3 /1 M Al(NO3)3 and the solution centrifuged at 3000

rpm for 10 minutes.

For thorium and uranium separation, Eichrom TEVA and UTEVA columns (2 mL pre-

packed columns, Triskem, France) were utilised. Firstly, a TEVA column was

preconditioned with 5 mL of 3 M HNO3 before the solution was loaded. The beaker was

rinsed with 5 mL of 3 M HNO3 and transferred onto the column. The thorium fraction

was retained on the column while the uranium fraction passed through. A further 30 mL

of 3 M HNO3 was passed through the column and the eluent was saved for uranium

purification. The thorium fraction was eluted with 20 mL 9 M HCl followed by 5 mL of

6 M HCl, and the eluate was taken to dryness under a heat lamp.

82

To purify the uranium, a UTEVA column was preconditioned with 3 M HNO3 before

the uranium solution was loaded. After passage of the sample, the column was

converted to the chloride form by adding 5 mL of 9 M HCl and 20 mL of 5 M HCl in

0.05 M H2C2O4. Finally, uranium was stripped with 15 ml of 1 M HCl and the solution

taken to dryness with a heat lamp.

For electrodeposition, 2.5 ml of 5 wt. % NaHSO4, 2 mL of deionised water and 5 mL of

15 wt. % NaHSO4 was added to the residue of purified uranium and thorium fraction

and heated. The solution was transferred to an electrodeposition cell and rinsed with 3

ml deionised water, and 1 ml of 20 g/L (NH4)2C2O4 plating solution was added. The

current was adjusted to 0.5 A for 5 minutes and then to 0.75 A for 90 minutes. 1 minute

before the end, 2 ml of 25 wt. % potassium hydroxide was added and the power was

turned off. The solution was discarded and the cell was washed with 2 ml 5 wt. %

NH4OH. Finally, the stainless-steel counting source was rinsed consecutively with small

volumes of ethanol and acetone.

3.4.5 Total radium

15-30 g amounts of dry sediments were sealed using insulating tape into double

polypropylene containers and put aside for at least four weeks, to avoid the escape of

222Rn and allow establishment of secular equilibrium between radium, radon and the

short-lived daughters, 214

Bi and 214

Pb. The total activity concentrations of the radium in

the sediments were measured by gamma spectrometry with a high purity germanium

(HPGe) detector and 45% relative efficiency at 1.33 MeV. Before measurement of the

samples, two standards were prepared by dispersing a known amount of 226

Ra

homogeneously through two samples, one with low organic content and one with high

organic content, which both had low radium contents. This allows compensation for the

effects of density and chemical composition. The samples were prepared in the same

physical geometries (height, volume and density) as the standard, since the sample and

83

the standard should have the same geometry, in order to make the calculation simple.

The method was checked using a standard reference material, stream sediment 314,

supplied by the International Atomic Energy Agency (IAEA).

The samples were counted for 12 hours, and the activity concentration of 226

Ra was

estimated from measurements of the 214

Bi gamma line at 609 keV and the 214

Pb gamma

line at 352 keV.

3.4.6 Radium separation

Radiochemical separation of radium was based on the method proposed by Smith and

Mercer (1970) using 225

Ra (150 mBq) as a radiotracer. After leaching the ashed

sediment as described for U/Th, the volume was made up to 50 ml using 0.1 M HNO3.

Radium was coprecipitated with PbSO4 after adding 1 mL of concentrated H2SO4, 2 g

K2SO4 and 1 ml of 0.24 M of Pb(NO3)2, consecutively. The solid was centrifuged in a

50 mL tube at 3000 rpm for 10 minutes, and then washed with 20 mL of a mixture of

0.2 M H2SO4/0.1 M K2SO4.

The precipitate was dissolved in 5 mL of 0.1 M ethylenediaminetetraacetic acid

(EDTA)/NH4OH (pH 10), passed through an anion exchange column (Bio-Rad AG1-

X8, 100-200 mesh, chloride form, 5 x 0.5 cm) to remove sulphate and washed with 13

mL 0.01 EDTA/NH4OH. To the eluate, 1 ml 5 M CH3COONH4 was added (pH 4.5) and

the solution was passed through a cation exchange column (Bio-Rad AG50W-X12,

200- 400 mesh, 8 x 0.7 cm) at a flow rate of 1 mL/minute. The column was previously

conditioned with 15 mL 1.5 M CH3COONH4 followed by 15 mL 0.25 M CH3COONH4.

50 mL 1.5 M CH3COONH4/0.1 M HNO3 was passed through this column to remove Pb

and Ac, while Ba was eluted by washing the column with 40 mL 2.5 M HCl. Finally, Ra

was eluted with 25 mL 6 M HNO3, and this solution was evaporated to dryness with

care. The radium was redissolved in an organic electrolyte solution (9 mL ethanol in 1

84

mL 0.1 M HNO3 and 2 mL 0.05 HCl) and electroplated on to a stainless steel planchette

at 120 mA for 30 minutes.

The prepared alpha sources were measured by alpha spectrometry (CANBERRA Model

7401) equipped with passivated implanted planar silicon (PIPS) detectors (Canberra,

Belgium, model A450) with 450 mm2 active area and alpha resolution (FWHM) 20 keV

at the 5.486 MeV alpha line. The planchettes were placed at ~5 mm distance from the

detector and a vacuum was applied. In these conditions, an absolute counting efficiency

of ~25% can be achieved. The acquisition time ranged from 1 to 10 days, depending on

the activity of the sample. Pulses were collected and spectral analyses were performed

using Genie 2000 3.1 software. Errors for individual measurements were estimated from

the measured counts and ranged from 3.3 to 15.8%.

3.4.7 Quality control

The analysis conducted, either for the total or the leached fraction, of the radionuclides

was tested by regular quality control methods. For the radiochemical separation, the

whole method was validated using blank samples spiked with the tracer, standard

additions and standard reference material (IAEA-314). The blank analyses always gave

less than 5 counts in each uranium region of interest, whereas all the sample analyses

are based on signals of at least 100 counts. In the standard additions, where a known

amount of 238

U was added to three duplicate samples and then the separation was

performed on the two samples, the measured uranium recoveries were 89 ± 13%, 116 ±

12% and 87 ± 15% of the added uranium. Aqua regia was used to leach the

radionuclides associated with organic matter and secondary minerals. The results for the

reference material were close to the recommended values (Table 3.2).

85


3.5.1 Characterisation of the stream sediments

Table 3.1 presents the mineralogical composition of the sediments from powder X-ray

diffraction measurements. It is obvious that the dominant mineral phase in all samples is

quartz, a primary mineral, with other minerals (albite, muscovite and kaolinite) also

present in the sediments. Sediments in streams coming from the north side of the valley

toward the River Noe comprise almost entirely quartz, while the majority of those

running from the south side of the valley toward the River Noe have accessory minerals

in addition to quartz.

Loss on ignition of the sediments ranged from 2.0 to 18.0 %, as can be seen in Table

3.1. The values can be related to the XRD results, with locations dominated by quartz

having low organic matter, while those with clays are also rich in organic matter.

3.5.2 238U,

234U,

230Th and

226Ra contents of the sediments

The activity concentrations of the total and the leached 238

U, 234

U, 230

Th and 226

Ra, as

well as 234

U/238

U and 230

Th/238

U isotopic ratios in Edale sediments are summarised in

Tables 3.3 and 3.4. The total 238

U activity concentrations range from 9.0 to 184.0 Bq kg-

1 and the leached from 5.0 to 91.0 Bq kg

-1. The total

234U activity concentrations are in

the range from 12.0 to 171.0 Bq kg-1

, and the leached from 5.0- 90.0 Bq kg-1

. The total

activity concentrations of the

230Th are from 9.0 to 200.0 Bq kg

-1 and the leached

fractions from 3.0 to 98.0 Bq kg-1

. The total 226

Ra activity concentrations are in the

range from 15.0 to 179.0 Bq kg-1

, and those from leaching are from 8.0 to 193.0 Bq kg-1

(although the highest leaching result is nominally higher than the highest result from

gamma spectrometry, the two results are essentially the same within measurement

error). The activity concentrations of the radionuclides extracted by leaching, relative to

total dissolution, are in the range of 30-70% for the uranium, 30-75% for the thorium

and 30-100% of the radium. This disparity of the radionuclide content may reflect

86

different source terms, as well as differing associations with secondary minerals and

organic matter content along the valley.

3.5.3 Fractionation of the radionuclides

3.5.3.1 (234U/

238U) activity ratios

For the total dissolution (Fig. 3.2), all of the sediments from Edale, except S8 and S23,

show 234

U/238

U > 1, within the error, and an average ratio of 1.1. For the acid leaching

(Fig. 3.3) the mean 234

U/238

U ratio is 1.2, more than the total dissolution results as might

be expected from the origins of disequilibrium.

These results cannot be explained by weathering processes only, since the expected

value of this ratio is less than unity in sediment and greater than unity in water (Pogge

von Strandmann et al., 2006; Vigier et al., 2001). However, similar results have been

published from previous studies of sediments and suspended matter from rivers in

Eastern England (Plater et al., 1992), in organic-rich sediments from rivers in Sweden

(Andersson et al., 1998) and in lowland rivers of the Amazon (Dosseto et al., 2006a).

The likely explanation is adsorption of a uranium component fraction with 234

U/238

U >1

from the water column onto the sediments’ mineral surfaces. It has been demonstrated

experimentally that some minerals such as clays or iron/manganese oxide phases are

efficient at removing radionuclides from the soluble phase (Duff et al., 2002). The

organic matter content may also contribute significantly to chemical fractionation

between uranium and its long-lived daughters (Dosseto et al., 2006b). Therefore, the

234U and

238U activities in Edale stream sediments may reflect the influence of both

reactive secondary minerals and organic matter.

3.5.3.2 (230Th/


If weathering processes primarily drive U-series fractionation, the 230

Th/238

U activity

ratio is expected to be > 1 in sediments and < 1 in river water (Plater et al., 1992; Vigier

et al., 2001). For the total dissolutions, only four samples (E1, E8, E10 and E22) show a

clear 230

Th/238

U > 1 (Fig. 3.4), while five samples of the leaching analyses (E1, E13,

87

E17, E24 and E25) display 230

Th/238

U > 1 (Fig. 3.5). This trend is not common, but

similar effects have been seen in samples of colloids and particulate matter from

lowland rivers of the Amazon basin, and the Kalix River in Sweden (Dosseto et al.,

2006a; Porcelli et al., 2001). Preferential complexation of thorium with dissolved

organic matter, which enhances its mobility, and hence leads to depletion in sediments

compared to uranium, may be the cause (Plater et al., 1992). As the disequilibrium

between 238

U and 234

U can be at least partly attributed to the organic matter content of

Edale valley sediments, a similar effect could be expected for the 230

Th/238

U activity

ratio.

3.5.3.3 (226Ra/


In normal conditions, mobility of Ra, Th and U has been reported to be in the order: U >

Ra > Th (Ivanovich and Harmon, 1992). From the total analyses and the acid leaching

results, given in Table 3.3 and Table 3.4 respectively, all of the streams in the valley,

except E8, E12 and E13, exhibit 226

Ra/238

U activity ratios > 1, indicating enrichment of

the daughter over the parent uranium. Elevated 226

Ra/238

U ratios, up to 9, have been

reported in organic-rich soil from Cronamuck Valley, Ireland (Dowdall and O'Dea,

2002). This is consistent with the expected pattern of mobility and may reflect efficient

adsorption of 226

Ra onto organic matter and mineral surfaces, comparable to the

possible oxidation and greater mobility of the uranium (Blanco et al., 2005). The

streams running near to the landslips (e.g. S3, S4, S5, S7 and S24) on the southern side

of the valley (Fig. 3.1) seem to show higher radium contents relative to the 238

U parent.

The 226

Ra/238

U disequilibrium was greater in the aqua regia leaches, compared with the

total dissolutions for all streams except S8, S12 and S13. The results suggest that, both

226Ra and

238U are subject to adsorption. However, adsorption on sediment is more

favourable for the divalent ion radium through precipitation and ion exchange with

calcium, so increasing 226

Ra content relative to 238

U (Lehto and Hou, 2010).

88

3.5.4 234U/

238U and

230Th/

238U isotopic ratio diagram

In order to examine the hypothesis that some Edale sediments do not conform to the

simple weathering processes example in U-series geochemical fractionation, plots of

(230

Th/238

U) against (234

U/238

U) activity ratios of the total and the leached fractions are

shown in Fig. 3.6 and Fig. 3.7. In simple leaching during weathering, it is not possible

to attain certain isotopic ratios in solid components such as sediments or suspended

matter, (Chabaux et al., 2008). These specific areas are identified as “Complex zones”

and have been coloured in grey, as in Figurer 3.6 and 3.7. However, plotting the data

obtained for both total dissolution and aqua regia leaching of the Edale sediments

revealed that the 230

Th/238

U and 234

U/238

U ratios fall in the complex zones. Within the

complex zone, some points from total dissolution plotted in the region representing

depletion in 234

U and 230

Th (Fig. 3.6), while majority of the aqua regia leach results

were in the region enriched in 234

U and 230

Th (Fig. 3.7). This is consistent with the idea

that radionuclides associated with the sediment fractions other than the residual are

more likely to be adsorbed from the water onto the sediment surfaces. Such complexed

behaviour has been observed previously (Dosseto et al., 2008) and, again, suggests the

need for alternative hypotheses to explain this behaviour other than the simple

weathering suggestion.

3.5.5 Hierarchical cluster analysis

Hierarchical cluster analysis (HCA) is a multivariate analysis technique designed to

categorize relatively similar samples, where there may not be a simple discriminator,

into separate groups. The technique has been recently used to identify associations

between sampling locations and a range of variables for river sediments (Gielar et al.,

2012; Guillén et al., 2012).

Prior to the calculations, all data were Z-scored, so that all variables carried equal

weight in the analysis. Z-scores measure the distance of the raw data from the mean of

that variable in terms of the standard deviation. The z-score (Zi) is given by,

89

||S

xxZi i

where x is the mean value, and S is the standard deviation of the whole population.

Following Z-scoring, all of the samples were plotted as points in 5-dimensional

hyperspace, where each (Z-scored) variable is used to define a coordinate, i.e., if a

sample has variable values of v, w, x, y and z, then the coordinate of the sample in

hyperspace will be (v,w,x,y,z).

The technique assumes that the difference between two samples increases with the

distance between the samples in hyperspace. The samples are clustered together, such

that, at each step, the sum of squares of the differences between the samples and their

cluster centres is minimised. In the first step, the two closest samples are linked. In the

next step, another sample can join the first cluster or two different samples can combine

to initiate a second, depending upon which action will minimise the sum of the square

of the distances. This procedure is repeated until, at the final stage, all of the samples

are linked together.

The similarity or distance between samples is measured by standard procedures, such as

Euclidean distance, which is an extension of Pythagoras’ theorem to a multidimensional

space. For the Euclidean distance, the square of the distances between two samples in a

multidimensional space is equal to the sum of squared differences of their coordinates

(Kim, 2000). The Euclidean metric is that used to measure the distance between points

in real space. However, other metrics are available, such as City block, where the

overall distance between points is the linear sum of the differences in their coordinates.

In this work, separate cluster analyses were performed with both distance metrics, since

if a result is real and significant, then it should be independent of the distance metric.

Mahalanobis distance is a statistical approach to identifying outliers within a set of

multivariable data. It measures the distance of a variable from the centroid

90

(multidimensional mean) of a distribution, given the covariance (multidimensional

variance) of the distribution (Kim, 2000).

For the total dissolution results, two outliers (E8 and E23) of the 25 sediment samples

from Edale valley were detected and excluded, since these outliers could not be

assigned to any of the four groups identified by the dendrogram, as shown in Fig. 3.8. In

total, 23 sediment samples were classified based on five variables: [238

U], [234

U],

[230

Th], [226

Ra] and loss on ignition. The groups were tested to examine any overlap

between them. The results revealed that the groups are separated clearly to five clusters.

The average concentrations of the samples in the five groups and their members are

given in Table 3.5. The samples clustered in group T1 (E3, E6, E9, E10, E13, E14, E15,

E20 and E21) represent samples with the lowest radionuclides and organic matter

contents. The samples in group T2 (E2, E5, E11, and E12) have a mean radionuclide

twice that of group T1, as well as a higher organic matter content. Group T3 (E1, E4

and E22) has radionuclide concentrations about three times the background level (T1)

although the organic matter content is similar to that of group T2. Group T4 (E19, E24

and E25) shows the highest organic matter content, but the concentrations of 238

U, 234

U

and 226

Ra are same as group T3. The locations in group T5 (E7, E16, E17 and E18) are

rich in organic matter, as group T4; however, the radionuclide concentrations are lower,

comparable to group T2.

For the results from aqua regia leaching, two outliers (again samples E8 and E23) of the

25 sediment samples from Edale Valley were excluded, since they could not be

assigned to any of the four groups identified by cluster analysis (Fig.3.9). In total, 23

sediment samples were classified based on the same five variables as before. The groups

were tested to examine any overlap between them but they are separated clearly,

although groups L1 and L3 are close.

91

The average concentrations of the samples in the four groups and their members are

given in Table 3.6. The samples clustered in group L1 (E3, E5, E6, E9, E10, E11, E13,

E14, E15, E20 and E21) represent samples with lower radionuclide contents, as well as

low organic matter content. Group L3 (E7, E16 and E17) has radionuclide

concentrations that are very similar to group L1, and these two groups have only

separated because of differences in organic matter content. The samples in group L2

(E1, E2, E12, and E18) show higher radionuclide concentrations, while Group L4 (E19,

E22, E24 and E25) shows the highest concentrations of 238

U, 234

U and 226

Ra, as well as

high organic matter content. The sample E4 was separated out as a single group,

because of its higher radium content. In addition, the outliers (E8 and E23) contain high

radionuclide concentrations and organic matter.

Comparing the cluster analyses results, from aqua regia leaching and total dissolution,

revealed that both outcomes eliminate the same two samples, E8 and E23, as outliers.

The obvious difference is that, an additional group is separated out from the total

dissolution results compared with those from aqua regia leaching.

The average values of the radionuclides of groups T1 (E3, E6, E9, E10, E13, E14, E15,

E20 and E21) and L1 (E3, E5, E6, E9, E10, E11, E13, E14, E15, E20 and E21), from

both aqua regia leaching and total dissolution data, represent background levels of the

radionuclides. Organic matter content is also low (4%) in both groups. This suggests

that, if the sample contains nothing much in the leach, it will contain nothing much in

the total.

The average content of radionuclides of groups L4 (E19, E22, E24 and E25) and T3

(E1, E4 and E22), from both aqua regia leaching and total dissolution data, show the

highest values among the groups. On average, group L4 shows high organic matter

(14%) compared with that of group T3 (6%). The average concentration of the

92

radionuclides in another group, T4 (E19, E24 and E25) is same as T3, with a difference

in organic content (OM 16%). The similarity in the average radionuclides in T3 and T4

suggests an association of the radionuclides with the primary minerals. However, the

radium content in T3, T4 and L4, despite the difference in the extraction method,

suggests a complex mechanism in binding/retaining radium.

On average, relatively low concentrations of radionuclides were observed in groups T5

(E7, E16, E17 and E18) and L3 (S7, S16 and S17), in association with high organic

matter (15% and 17% respectively). However, while organic matter content appears to

bear some relationship to radionuclide content, it is not a good discriminator, as

illustrated by groups L2 (E1, E2, E12 and E18) and T2 (E2, E5, E11 and E12), where

both groups have a relatively low OM content (7% and 5% respectively) but

nevertheless have high radionuclide activities.

Comparison of the cluster analysis results, for both aqua regia leaching and total

dissolution data, with the sampling locations in Fig. 3.1 shows that none of the groups

identified by cluster analysis has all of its members gathered in one area of the valley.

However, samples with the highest radionuclide concentrations from aqua regia

leaching (E19, E22, E24 and E25) are located to the south of the river, and three of them

are adjacent, in the south-east of the sampling area.

For the total dissolution results, groups T3 and T4 (which represent, relatively, the

higher activity samples) are also concentrated on the southern side of the valley,

generally close to landslips. In particular, the samples with higher radium values are

located in streams that are likely to receive runoff water from the slips. It is possible that

the landslips have exposed relatively fresh and/or uranium-rich material higher up the

valley sides, and that radium is leaching from that material and being transported down

the streams, where it is sorbing onto the stream sediments. This would also explain the

93

234U/

238U ratios > 1, since

234U would be expected to leach preferentially from the

material exposed by the landslips, leading to higher sorbed concentrations in the stream

sediments.

3.6 Conclusions

U-series isotopes have been measured in stream sediments, applying two dissolution

methods, in an effort to understand chemical weathering and physical erosion in the

Edale Valley. This study showed considerable variability in radionuclide

concentrations, even over a fairly small geographical area, albeit a geologically complex

one. This variation suggests an interplay of the parent materials, organic matter and

secondary minerals in the sediments. The daughter/parent isotopic ratios revealed

complex U-series disequilibria. Adsorption of uranium onto mineral surfaces and/or

organic matter and migration of thorium complexed by organic matter are likely to be

major impacts on these disequilibria. Plots of (234

U/ 238

U) against (230

Th/ 238

U) indicate

that weathering processes in the Edale Valley are not simple. Cluster analysis provides

insight into radionuclide behaviour and suggests a relationship between the landslips in

the Noe Valley and the stream sediment isotope concentrations. It is possible that

uranium-containing material has been exposed by the slips, and that 226

Ra and 234

U are

being released into the runoff water, and then becoming sorbed onto the stream

sediments.

Acknowledgements

The authors appreciate the financial fund from the Islamic Development Bank (IDB),

Jeddah, Saudi Arabia and are also grateful to the UK Natural Environment Research

Council (NERC) for support.

94

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Buckau, G., 2012. Real system analyses/natural analogues. Applied Geochemistry 27,

490-500.

Peacock, J.D., Taylor, K., 1966. Uraniferous Collophane in the Carboniferous

Limestone of Derbyshire and Yorkshire. NERC Bulletin of the Geological Survey of

Great Britain No. 25 London.


U / 238


technique for studying uranium mobility in the Opalinus Clay at Mont Terri,

Switzerland. Applied Radiation and Isotopes 68, 984-992.

Plater, A.J., Ivanovich, M., Dugdale, R.E., 1992. Uranium series disequilibrium in river

sediments and waters: the significance of anomalous activity ratios. Applied

Geochemistry 7, 101-110.

Pogge von Strandmann, P.A.E., Burton, K.W., James, R.H., van Calsteren, P., Gislason,

S.R., Mokadem, F., 2006. Riverine behaviour of uranium and lithium isotopes in an

actively glaciated basaltic terrain. Earth and Planetary Science Letters 251, 134-147.

Pogge von Strandmann, P.A.E., Burton, K.W., Porcelli, D., James, R.H., van Calsteren,

P., Gislason, S.R., 2011. Transport and exchange of U-series nuclides between

suspended material, dissolved load and colloids in rivers draining basaltic terrains. Earth

and Planetary Science Letters 301, 125-136.

97

Porcelli, D., Andersson, P.S., Baskaran, M., Wasserburg, G.J., 2001. Transport of U-

And Th-series nuclides in a Baltic Shield watershed and the Baltic Sea. Geochimica et





Suresh, G., Ramasamy, V., Meenakshisundaram, V., Venkatachalapathy, R.,

Ponnusamy, V., 2011. A relationship between the natural radioactivity and

mineralogical composition of the Ponnaiyar river sediments, India. Journal of


Vigier, N., Bourdon, B., Turner, S., Allegre, C.J., 2001. Erosion timescales derived

from U-decay series measurements in rivers. Earth and Planetary Science Letters 193,

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Vigier, N., Burton, K.W., Gislason, S.R., Rogers, N.W., Duchene, S., Thomas, L.,

Hodge, E., Schaefer, B., 2006. The relationship between riverine U-series disequilibria

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shallow water sediments in the upper Carboniferous of northern England. Journal of

Sedimentary Research 36, 90-114.

98

Table 3.1 Edale sediment sample coordinates, loss on ignition and mineralogy

Sample ID Coordinates Loss on ignition % Mineralogy

E1 53⁰ 21.558' N; 1⁰ 50.107' W 5 Q, M, A

E2 53⁰ 21.549' N; 1⁰ 50.042' W 7 Q, M, A

E3 53⁰ 21.526' N; 1⁰ 49.722' W 4 Q, M, A

E4 53⁰ 21.518' N; 1⁰ 49.561' W 6 Q, M, A

E5 53⁰ 21.513' N; 1⁰ 49.567' W 6 Q, M, A

E6 53⁰ 21.518' N; 1⁰ 49.276' W 5 Q, M, A, K

E7 53⁰ 21.508' N; 1⁰ 49.187' W 14 Q

E8 53⁰ 21.475' N; 1⁰ 48.903' W 9 Q, M, K, A

E9 53⁰ 21.952' N; 1⁰ 49.350' W 5 Q, A

E10 53⁰ 21.895' N; 1⁰ 49.458' W 6 Q, A,

E11 53⁰ 21.730' N; 1⁰ 49.711' W 5 Q, A

E12 53⁰ 21.696' N; 1⁰ 49.840' W 3 Q, A

E13 53⁰ 22.108' N; 1⁰ 48.893' W 3 Q, A

E14 53⁰ 22.221' N; 1⁰ 48.488' W 3 Q, A

E15 53⁰ 22.178' N; 1⁰ 48.350' W 3 Q, A

E16 53⁰ 22.239' N; 1⁰ 47.947' W 18 Q, A

E17 53⁰ 22.211' N; 1⁰ 48.009' W 17 Q, A

E18 53⁰ 21.563' N; 1⁰ 50.471' W 11 Q, A

E19 53⁰ 21.648' N; 1⁰ 50.660' W 19 Q, A

E20 53⁰ 21.882' N; 1⁰ 50.842' W 2 Q, A

E21 53⁰ 21.875' N; 1⁰ 50.912' W 4 Q, K, A

E22 53⁰ 21.769' N; 1⁰ 48.406' W 6 Q, A

E23 53⁰ 21.735' N; 1⁰ 48.678' W 11 Q, A

E24 53⁰ 21.783' N; 1⁰ 48.303' W 15 Q

E25 53⁰ 21.798' N; 1⁰ 48.237' W 15 Q, M, A

(Q= Quartz, M= Muscovite, K= Kaolinite, A= Albite)

99


Ra and 238

U in

IAEA-314 stream sediment reference material

226

Ra Bq.kg-1

238

U mg.kg-1

Measured 774 ± 24 58 ± 1.1

Recommended 732 56.8

95% Confidence interval 678 – 787 52.9 – 60.7

Value from leaching using aqua regia 490 ± 1.5 43 ± 1.6

% of the leached fraction in the recommended value 67% 76%

100


dry weight) of the total 238

U, 234

U, 230

Th,

226Ra and

234U/

238U,

230Th/

238U,

226Ra/


Valley (± 1σ counting statistics uncertainties)

ID 238

U 234

U 230

Th 226

Ra 234

U/238

U 230

Th/238

U 226

Ra/238

U

E1 73 ± 4 75 ± 4 30 ± 4 71 ± 2 1.03 ± 0.07 0.42 ± 0.15 0.97 ± 0.07

E2 43 ± 4 49 ± 4 46 ± 5 53 ± 2 1.15 ± 0.14 1.09 ± 0.36 1.26 ± 0.20

E3 24 ± 3 24 ± 3 20 ± 4 45 ± 2 1.01 ± 0.16 0.82 ± 0.37 1.88 ± 0.19

E4 30 ± 3 32 ± 3 31 ± 5 131 ± 15 1.06 ± 0.16 1.04 ± 0.44 4.33 ± 0.69

E5 34 ± 3 38 ± 3 34 ± 4 66 ± 3 1.10 ± 0.14 0.98 ± 0.36 1.93 ± 0.13

E6 24 ± 3 27 ± 3 19 ± 4 29 ± 1 1.12 ± 0.17 0.78 ± 0.37 1.19 ± 0.17

E7 25 ± 3 29 ± 3 19 ± 4 57 ± 3 1.18 ± 0.17 0.77 ± 0.35 2.31 ± 0.11

E8 81 ± 5 72 ± 4 56 ± 6 86 ± 3 0.89 ± 0.07 0.69 ± 0.24 1.05 ± 0.07

E9 22 ± 3 24 ± 3 45 ± 4 36 ± 1 1.05 ± 0.20 2.01 ± 0.68 1.61 ± 0.16

E10 36 ± 4 41 ± 4 20 ± 4 39 ± 2 1.13 ± 0.16 0.54 ± 0.24 1.07 ± 0.09

E11 46 ± 4 45 ± 4 37 ± 5 70 ± 4 0.97 ± 0.12 0.81 ± 0.32 1.52 ± 0.09

E12 35 ± 3 38 ± 3 33 ± 5 61 ± 3 1.09 ± 0.14 0.94 ± 0.38 1.75 ± 0.10

E13 19 ± 2 21 ± 2 21 ± 4 15 ± 1 1.10 ± 0.17 1.06 ± 0.48 0.78 ± 0.19

E14 18 ± 2 18 ± 2 14 ± 3 29 ± 1 1.03 ± 0.19 0.78 ± 0.39 1.68 ± 0.16

E15 9 ± 2 12 ± 2 9 ± 3 18 ± 1 1.22 ± 0.28 0.97 ± 0.54 1.86 ± 0.25

E16 24 ± 3 23 ± 2 25 ± 5 40 ± 2 0.97 ± 0.15 1.02 ± 0.47 1.65 ± 0.12

E17 41 ± 3 53 ± 4 38 ± 6 44 ± 2 1.28 ± 0.14 0.92 ± 0.36 1.06 ± 0.14

E18 33 ± 3 47 ± 4 27 ± 4 48 ± 2 1.41 ± 0.18 0.81 ± 0.33 1.45 ± 0.14

E19 64 ± 4 59 ± 4 33 ± 4 107 ± 4 0.93 ± 0.08 0.52 ± 0.19 1.68 ± 0.14

E20 14 ± 2 15 ± 2 19 ± 4 36 ± 2 1.09 ± 0.20 1.34 ± 0.64 2.53 ± 0.21

E21 20 ± 2 24 ± 3 20 ± 4 33 ± 2 1.20 ± 0.19 0.98 ± 0.45 1.63 ± 0.28

E22 86 ± 5 84 ± 5 64 ± 7 119 ± 4 0.98 ± 0.08 0.74 ± 0.26 1.38 ± 0.10

E23 184± 8 170± 8 200 ± 13 179 ± 8 0.93 ± 0.06 1.09 ± 0.28 0.97 ± 0.06

E24 48 ± 4 65 ± 4 56 ± 7 104 ± 4 1.37 ± 0.14 1.17 ± 0.41 2.17 ± 0.22

E25 51 ± 4 63 ± 4 38 ± 5 89 ± 4 1.24 ± 0.12 0.75 ± 0.29 1.76 ± 0.17

101


dry weight) of the leached 238

U, 234

U, 230

Th,

226Ra and

234U/

238U,

230Th/

238U,

226Ra/


Valley (± 1σ counting statistics uncertainties)

ID 238

U 234

U 230

Th 226

Ra 234

U/238

U 230

Th/238

U 226

Ra/238

U

E1 25 ± 2 31 ± 3 18 ± 2 56 ± 2 1.20 ± 0.15 0.70 ± 0.23 2.19 ± 0.18

E2 19 ± 1 25 ± 2 28 ± 3 53 ± 4 1.32 ± 0.14 1.47 ± 0.46 2.81 ± 0.02

E3 7 ± 1 9 ± 1 6 ± 1 17 ± 1 1.33 ± 0.25 0.93 ± 0.40 2.53 ± 0.38

E4 12 ± 1 14 ± 1 12 ± 1 180 ± 12 1.17 ± 0.09 1.03 ± 0.27 15.61 ±0.81

E5 14 ± 1 18 ± 1 19 ± 1 21 ± 1 1.29 ± 0.13 1.33 ± 0.35 1.51 ± 0.10

E6 7 ± 1 8 ± 1 7 ± 1 13 ± 1 1.26 ± 0.25 1.05 ± 0.43 2.04 ± 0.32

E7 7 ± 1 8 ± 1 6 ± 1 14 ± 1 1.16 ± 0.23 0.95 ± 0.41 2.18 ± 0.33

E8 50 ± 6 49 ± 6 43 ± 6 41 ± 2 0.98 ± 0.16 0.84 ± 0.32 0.81 ± 0.09

E9 10 ±1 12 ± 1 13 ± 1 14 ± 1 1.16 ± 0.09 1.27 ± 0.34 1.34 ± 0.07

E10 17 ± 1 20 ± 1 20 ± 1 19 ± 1 1.15 ± 0.12 1.16 ± 0.27 1.07 ± 0.07

E11 9 ± 1 10 ± 1 19 ± 2 24 ± 1 1.13 ± 0.15 2.07 ± 0.64 2.62 ± 0.24

E12 25 ± 2 27 ± 1 40 ± 3 20 ± 1 1.09 ± 0.13 1.63 ± 0.43 0.80 ± 0.07

E13 18 ± 1 22 ± 1 6 ± 1 9 ± 1 1.19 ± 0.12 0.30 ± 0.13 0.51 ± 0.03

E14 5 ± 1 6 ± 1 5 ± 1 10 ± 1 1.12 ± 0.30 1.08 ± 0.52 2.10 ± 0.43

E15 5 ± 1 5 ± 1 3 ± 1 8 ± 1 1.02 ± 0.32 0.68 ± 0.42 1.73 ± 0.39

E16 7 ± 1 9 ± 1 6 ± 1 9 ± 1 1.23 ± 0.22 0.83 ± 0.36 1.31 ± 0.19

E17 13 ± 1 15 ± 1 9 ± 1 41 ± 2 1.16 ± 0.10 0.72 ± 0.18 3.22 ± 0.19

E18 17 ± 1 20 ± 1 15 ± 2 39 ± 1 1.21 ± 0.12 0.89 ± 0.29 2.36 ± 0.16

E19 32 ± 3 33 ± 3 30 ± 3 86 ± 6 1.04 ± 0.14 0.96 ± 0.32 2.72 ± 0.25

E20 8 ± 1 10 ± 1 14 ± 1 11 ± 1 1.23 ± 0.17 1.83 ± 0.54 1.43 ± 0.18

E21 11 ± 1 15 ± 1 14 ± 1 35 ± 1 1.36 ± 0.11 1.27 ± 0.34 3.10 ± 0.16

E22 40 ± 1 47 ±15 42 ± 7 84 ± 6 1.17 ± 0.18 1.05 ± 0.46 2.09 ± 0.22

E23 91 ±14 90 ±14 98 ±19 193 ± 10 0.99 ± 0.22 1.07 ± 0.50 2.12 ± 0.33

E24 29 ± 3 39 ± 4 22 ± 1 115 ± 6 1.36 ± 0.19 0.77 ± 0.16 3.98 ± 0.27

E25 34 ± 3 47 ± 5 24 ± 1 92 ± 5 1.39 ± 0.21 0.70 ± 0.14 2.73 ± 0.27

102



(total dissolution) and loss on ignition (wt.%) of the hierarchical cluster analysis (S.D. =

standard deviation).

Group T1

238U

234U

230Th

226Ra L.O.I

E3 24 24 20 45 5

E6 24 27 19 29 6

E9 22 24 45 36 5

E10 36 41 20 39 6

E13 19 21 21 15 3

E14 18 18 13 29 2

E15 9 12 9 18 4

E20 14 15 19 36 2

E21 20 24 20 33 4

Mean 21 23 21 31 4

S.D. 8 8 10 10 2

Group T2

238U

234U

230Th

226Ra L.O.I

E2 43 49 46 53 7

E5 34 38 34 66 6

E11 46 45 37 70 5

E12 35 38 33 61 3

Mean 40 42 38 63 5

S.D. 6 5 6 7 2

103

Group T3

238U

234U

230Th

226Ra L.O.I

E1 72 75 30 70 6

E4 30 32 31 131 5

E22 86 84 64 119 6

Mean 63 64 42 106 6

S.D. 29 28 19 32 0.6

Group T4

238U

234U

230Th

226Ra L.O.I

E19 64 59 33 107 19

E24 48 65 56 104 15

E25 51 63 38 89 15

Mean 54 62 42 100 16

S.D. 9 3 12 10 2

Group T5

238U

234U

230Th

226Ra L.O.I

E7 25 29 19 57 13

E16 24 23 25 40 18

E17 41 53 38 44 18

E18 33 47 27 48 11

Mean 31 38 27 47 15

S.D. 8 14 8 7 4

104



(leached) and loss on ignition (wt.%) of the hierarchical cluster analysis (S.D. =

standard deviation).

Group L1

238U

234U

230Th

226Ra L.O.I

E3 7 9 6 17 5

E5 14 18 19 21 6

E6 7 8 7 13 6

E9 10 13 13 14 5

E10 17 20 20 19 6

E11 9 10 19 24 5

E13 18 22 6 9 3

E14 5 6 5 10 2

E15 5 5 3 8 4

E20 8 10 14 11 2

E21 11 15 14 35 4

Mean 10 12 12 16 4

S.D. 5 6 6 8 2

Group L2

238U

234U

230Th

226Ra L.O.I

E1 25 31 18 56 6

E2 19 25 28 53 7

E12 25 27 40 20 3

E18 17 20 15 39 11

Mean 21 26 25 42 7

S.D. 4 4 12 16 3

105

Group L3

238U

234U

230Th

226Ra L.O.I

E7 7 8 6 14 13

E16 7 9 6 9 18

E17 13 15 9 41 18

Mean 9 11 7 22 17

S.D. 4 4 2 17 3

Group L4

238U

234U

230Th

226Ra L.O.I

E19 32 33 30 86 19

E22 40 47 42 84 6

E24 29 39 22 115 15

E25 34 47 24 92 15

Mean 34 41 30 94 14

S.D. 5 7 9 15 5

106

Figure 3.1 Edale Valley, Derbyshire and the sampling points

107

Figure 3.2 234

U/238


Edale Valley

0.00

0.20

0.40

0.60

0.80

1.00

1.20

1.40

1.60

1.80

E1

E2

E3

E4

E5

E6

E7

E8

E9

E10

E11

E12

E13

E14

E15

E16

E17

E18

E19

E20

E21

E22

E23

E24

E25

23

4U

/23

8U

act

ivit

y r

atio

of

tota

l dis

solu

tio

n

Sample location

108

Figure 3.3 234

U/238

U activity ratios from aqua regia leaching analyses of sediments from

Edale Valley

0.00

0.20

0.40

0.60

0.80

1.00

1.20

1.40

1.60

1.80

E1

E2

E3

E4

E5

E6

E7

E8

E9

E10

E11

E12

E13

E14

E15

E16

E17

E18

E19

E20

E21

E22

E23

E24

E25

23

4U

/23

8U

act

ivit

y r

atio

of

aqua

regia

lea

chin

g

Sample location

109

Figure 3.4 230

Th/238


Edale Valley

0.00

0.50

1.00

1.50

2.00

2.50

3.00

E1

E2

E3

E4

E5

E6

E7

E8

E9

E10

E11

E12

E13

E14

E15

E16

E17

E18

E19

E20

E21

E22

E23

E24

E25

23

0T

h/2

38U

act

ivit

y r

atio

of

tota

l dis

solu

tio

n

Sample location

110

Figure 3.5 230

Th/238

U activity ratios from aqua regia leaching of sediments from Edale

Valley

0.00

0.50

1.00

1.50

2.00

2.50

3.00

E1

E2

E3

E4

E5

E6

E7

E8

E9

E10

E11

E12

E13

E14

E15

E16

E17

E18

E19

E20

E21

E22

E23

E24

E25

23

0T

h/2

38U

act

ivit

y r

atio

of

aqu

a re

gia

lea

chin

g

Sample location

111

Figure 3.6 234

U/238

U vs 230

Th/238

U diagram for total dissolution analyses of sediments

from Edale Valley (Grey colour represents complex zones)

112

Figure 3.7 234

U/238

U vs 230

Th/238

U diagram for aqua regia leaching of sediments from

Edale Valley (Grey colour represents complex zones)

113

Figure 3.8 Dendrogram illustrating cluster analysis, from total dissolution data, of


U], [234

U], [230

Th], [226

Ra] and

loss on ignition

114

Figure 3.9 Dendrogram illustrating cluster analysis, from aqua regia leaching, of


U], [234

U], [230

Th], [226

Ra] and

loss on ignition

115

Chapter Four

Geochemical characterisation of uranium and radium in

sediments near an abandoned uranium mine, Cornwall, UK

The material in the following section was given as an oral presentation at The American

Chemical Society (ACS), Spring Meeting, 07 - 11 April, 2013 in New Orleans, US. The

final draft was prepared for submission to Applied Geochemistry.

The author was involved in the collection of samples, performed the radiochemical

measurements, analysed the data, interpreted the results and wrote first draft and the

final version of the manuscript.

ICP-OES and ICP-MS were conducted by Mr Paul Lythgoe. IC analysis was conducted

by Mr Alastair Bewsher. Total dissolution of sediments was achieved by Mrs Cath

Davies. SEM was carried out under supervision of Dr John Waters. EMPA analysis

was carried out by Dr John Charnock. All are from School of Earth, Atmospheric and

Environmental Sciences. The University of Manchester.

116

Geochemical characterisation of uranium and radium in

sediments near an abandoned uranium mine, Cornwall, UK

Siddeeg, S.E., Bryan, N.D., and Livens, F.R.



Abstract

Water and sediment samples were taken from the vicinity of the abandoned South

Terras uranium mine in south-west UK and analysed for uranium and radium to explore

their geochemical dispersion. The radioactivity concentrations of the sediment samples

were measured using alpha spectrometry for uranium, and gamma spectrometry for

radium. Sequential chemical extraction was applied to selected sediments in order to

investigate the speciation of the radionuclides and their association with stable

elements. The activity ratio between uranium isotopes was used to explore the mobility

of uranium. Spectroscopic methods, scanning electron microscopy (SEM) and an

electron microprobe analyser (EMPA) were used to characterise the sediments. The

radiochemical results identified two samples with enhanced radioactivity. The

geochemical distribution of the radionuclides in these samples varies with the

operationally-defined fractions. The majority of the uranium was released from the

carbonate fraction but, in one sample, there was also a significant organic association of

uranium, while in the second sample, there was an association with the resistant

fraction. Geochemical distributions of the stable elements were different in both

samples. The activity ratio of 234

U/238

U shows different trends in the two sediments,

signifying the impact of organic matter and/or the exchange between water and

sediment. SEM and EMPA analysis identified uranium-bearing minerals in association

with potassium, calcium, iron, manganese and arsenic.

117

4 Introduction

Natural decay series radionuclides show elevated concentrations in all igneous rocks

(Plater et al., 1992). Therefore, regions with acid igneous rocks, such as granite, are

considered potential uranium mining areas. Following mining and milling processes,

large amounts of radioactive waste from the uranium decay chain are produced,

incorporated in spoil heaps and mill tailings (Carvalho et al., 2007). The tailings left

behind hold most of the radium in the ore, as a co-precipitate with barium or lead,

usually associated with fine grained material in the waste. Alpha emitters, such as 238

U,

234U,

230Th and

226Ra, are the most important isotopes in the radiological assessment of

many former uranium mining locations worldwide (Blanco et al., 2005; Carvalho and

Oliveira, 2007; Hancock et al., 2006; Marko Strok, 2010). Accordingly, continuous

radiological surveillance of these sites is required, even after cessation, to monitor

radioactivity in waste piles and spoil heaps.

The south-west of the UK, with its scattered granitic intrusions, has a rich history of

mining activities (Gillmore et al., 2001). In the context of radioactive deposits, the most

significant mine in Cornwall is that at South Terras (50⁰ 20.048' N 4⁰ 54.311' W),

which was the only UK mine worked primarily for uranium and, subsequently, radium.

The potential for uranium contamination of water and sediment collected from the River

Fal, which flows close to the vicinity of the mine, has been studied (Moliner-Martinez

et al., 2004). The results suggested that the uranium mine and spoil heaps at the mine

were not significant sources of uranium in the river water. This was supported by a

correlation found between the total cation concentrations and uranium in the surface

water, thereby suggesting that uranium in the river water originated from rock

weathering. Nevertheless, the concentration of uranium locally in sediment beneath the

outflow pipe was extremely high and reached up to 1000 ppm.

118

Nearly a century after the cessation of mining activities in South Terras, the spoil heaps

can be used as a natural laboratory in which to study the environmental geochemistry of

the natural radionuclides. The area is of particular interest due to the retention of the

long-lived radionuclides and the expected presence of a variety of heavy metals and

minerals (Gillmore et al., 2001). Therefore, surface water and surface sediments in the

River Fal are of interest from a radiological and toxicological perspective.

The aim of this study is to use the South Terras mine as a natural analogue in which to

explore the behaviour of the natural radionuclides from the U-decay series, mainly 238

U,

234U and

226Ra, associated with the uranium mining activities. It is proposed that this

will be achieved by:

Quantifying the distribution of natural radionuclides associated with uranium

mining, namely 238

U, 234

U and 226

Ra.

Characterising sediments with elevated radionuclide contents to investigate

the geochemical associations and mobility of the radionuclides.

Identifying factors affecting this transport, such as adsorption to organic

matter, association with different geochemical phases within the sediments,

relationship with trace elements and the role of physico-chemical parameters

of the river water (e.g. pH, Eh, total dissolved solids (TDS), anions and

cations in water).

4.1 The study area and sampling

In the south-west region of the UK, Cornwall’s high-temperature (300-500°C) veins,

oriented NE-SW and associated with diverse and complex mineralisation, have been

exploited for different elements, including copper, tin, iron and lead. Other low-

temperature veins (100- 300°C) cross the high-temperature veins. These contain a small

amount of pitchblende, and have been explored for cobalt, nickel, iron, lead, uranium

119

and then for radium (Purvis et al., 2004). Uranium was excavated mainly from

pitchblende (U3O8) and uraninite (UO2) as primary ores, but secondary minerals, such

as autunite [Ca(UO2)2(PO4)2.10H2O], zippeite [(UO2)3(SO4)2(OH)2.8H2O] and

torbernite [Cu(UO2)2(PO4)2.8H2O], were reported to be common in the area (Purvis et

al., 2004). The area around the abandoned mine is now covered by vegetation and the

heaps of waste materials have been reshaped by erosion.

Twenty locations along an approximately 2 km stretch of the valley of the River Fal,

running south from the South Terras mine site, were sampled for water and surface

sediments (Fig. 4.1). The water samples were collected in polyethylene bottles. As soon

as possible after collection (always within 12 hours), each water sample was divided

into three subsamples: unacidified, unfiltered (for physicochemical analysis, such as pH

and electrical conductivity); acidified, filtered (for elemental analysis); and unacidified,

filtered (for anions measurement). The filtration was conducted using 0.22 µm cellulose

acetate filters and the the samples were acidified with nitric acid (1 ml concentrated

HNO3 per 100 ml of water). In the field, the samples’ pH was measured using a pH-

meter (SevenEasy, Mettler-Toledo GmbH). The sediment samples were saved in Kraft®

paper envelopes. In the laboratory, the sediments were wet-sieved through 2 mm mesh

and left to air dry on open trays for several days. The dry sediments were disaggregated

gently using a mortar and pestle, and stored in plastic bottles.

4.2 Methodology

4.2.1 Physicochemical analysis of water

A pH meter, SevenEasy, Mettler-Toledo GmbH, and a probe were used to measure the

pH of the water samples from the River Fal in Cornwall and a Jenway 4010

conductivity meter with a probe was used for measuring of the specific conductivity

(S/cm).

120

The filtered water samples were measured in an ion chromatograph (IC) for chloride,

nitrate, and sulphate. The IC instrument consisted of a Metrohm 761 Compact ion

chromatograph, fitted with a Dionex Ion-Pac AG9-HC (guard), a Dionex Ion-Pac AS9-

HC analytical column and a conductivity detector. The backpressure was 2000 psi, the

mobile phase was 9 mM Na2CO3 and the eluent flow rate was 1.4 mL/min. A set of

standard solutions, with concentrations of 0.5, 3.0, 10.0 and 30.0 mg/L for chloride,

nitrate and sulphate was used for calibration. The detection limit was approximately

0.05 mg/L for most analytes.

4.2.2 Physicochemical properties of sediments

4.2.2.1 Loss on Ignition

Organic matter (OM) content was estimated from loss on ignition (Sutherland, 1998). A

porcelain crucible was ignited at 550 ⁰C for 30 minutes in a muffle furnace, then

allowed to cool in a desiccator and accurately weighed. From the bulk dry sediments,

1.0-2.0 g was placed in the crucible and weighed accurately, then transferred to a muffle

furnace and heated to 550 ⁰ C for 5 hours. The hot crucible, containing the residue, was

placed in the desiccator and cooled to ambient temperature. The crucible containing the

ashed sediment was weighed accurately and the loss on ignition as a percentage was

calculated.

4.2.2.2 Mineral identification using X-ray diffraction

The dry sediments were sieved through 80 mesh and a suitable amount (~0.5 g) of each

sample was placed on the sample holder. A smooth, flat surface was obtained using a

glass slide, before placing the sample in the specimen position of the XRD. Mineral

identifications were made using a Bruker D8Advance Powder diffractometer. The X-ray

is generated from a Cu X-ray tube (Kα with a wavelength of 0.152 nm, current 30 mA

at 40 kV) and the instrument is equipped with a standard scintillation detector. The

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scanning angle for the specimen was set from 5⁰ to 75⁰ with a step size of 0.02⁰/s and

an exposure time of 30 minutes. Phase identification was performed using Eva 14,

Bruker version 2008 pattern analysis tool.

4.2.3 Radioactivity content in sediments

4.2.3.1 Sediment dissolution

For total dissolution of the sediments, 0.2 g of the ashed sediment was placed in a

closed vessel and wetted overnight with a mixture of 1.0 mL deionised water, 3.0 mL

concentrated nitric acid and 6.0 mL concentrated hydrofluoric acid. The sample was

then digested in a microwave oven with ramping time 10 minutes to 140 ⁰C (~150 psi)

and 50 minutes holding time, and this was repeated three times before evaporation.

Finally, 2.0 mL of 20 % nitric acid was added to the residue and the volume was made

up to 20 mL with deionised water.

4.2.3.2 Uranium separation

The uranium separation was based on extraction chromatography methods (Carter et al.,

1999; Eichrom Technologies 2001). For total dissolution, approximately 40 mBq of

232U tracer was added to a suitable aliquot (20 to 22 mL) of solution and the solution

brought to near dryness under a heat lamp. Then, 5.0 mL conc. HNO3 was added to the

residue and the solution brought to near dryness under a heat lamp. The residue was

dissolved with 10.0 mL of 3.0 M HNO3/1 M Al(NO3)3 and the resultant solution was

centrifuged at 3000 rpm (about 6500 g) for 10 minutes.

An extraction chromatography column (UTEVA, 2.0 mL pre-packed column; Eichrom

resin, Triskem, France) was preconditioned with 5 ml 3.0 M HNO3 before loading the

solution. The beaker was washed with 5.0 mL 3.0 M HNO3 and the wash was passed

through the column. Then the column was rinsed with consecutive additions of 5.0 mL

of 3.0 M HNO3, 5.0 mL of 9.0 M HCl and 20.0 mL of 5.0 M HCl in 0.05 M H2C2O4.

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All these eluates were discarded and, finally, uranium was stripped with 15.0 mL of 1.0

M HCl. The eluent was evaporated to near dryness for electrodeposition in the presence

of 1.0 mL 10% (w/v) KHSO4 using a heat lamp.

For U electrodeposition, 2.5 ml of 5 wt. % NaHSO4, 2.0 ml of deionised water and 5.0

ml of 15.0 wt. % Na2SO4 were added to the residue of the purified U fractions and

heated gently until the residue dissolved. The solution was transferred to an

electrodeposition cell and rinsed in with 3.0 ml deionised water, then 1.0 ml of 20.0 g/L

ammonium oxalate plating solution was added. The current was adjusted to 0.5 A for 5

minutes and then to 0.75 A for 90 minutes. One minute before the end, 2.0 ml of 25.0

wt. % potassium hydroxide was added and the power was turned off. The solution was

discarded and the cell was washed with 2.0 ml 5.0 wt. % ammonium hydroxide.

Finally, the stainless-steel disk was rinsed consecutively with a small volume of 5.0 wt.

% ammonium hydroxide, ethanol and acetone before being dried on a hotplate at 200 °C

for 5 minutes.

The whole radiochemical separation method was validated using blank samples

(deionised water spiked with the tracer) and the IAEA RM-314 reference material. The

method was also tested using standard additions, in which a known amount of 238

U was

added to three duplicate samples and then the separation was performed.

4.2.4 Total radium in sediments

Total radium in the sediments was measured using gamma spectrometry with a high

purity germanium (HPGe) detector. Sample preparation is relatively straightforward. To

avoid the escape of 222

Rn gas, the samples were sealed in a double polypropylene

container and put aside for four weeks to reach secular equilibrium between 226

Ra,

222Rn,

214Bi and

214Pb, where activity concentrations of all these radionuclides will be

123

equal. The samples were then counted for 12 hours, and the activity concentration of

226Ra was estimated from measurements of the

214Bi gamma line at 609 keV and the

214Pb gamma line at 352 keV.

4.2.4.1 Quality control

The analysis conducted, either for the total or the leached fraction, of the radionuclides

was tested by regular quality control methods. As described above, the radiochemical

separation was validated using blank solutions spiked with the tracer, standard additions

and a standard reference material (IAEA-314). The blank analyses always gave less

than 5 counts in each uranium region of interest, whereas all the sample analyses are

based on signals of at least 100 counts. In the standard additions, where a known

amount of 238

U was added to three duplicate samples and then the separation was

performed on the two samples, the measured uranium recoveries were 92 ±12%, 116 ±

17% and 87 ± 11% of the added uranium. The results for the reference material were

close to the recommended values as can be seen in Table 4.3.

4.2.4.2 Sequential chemical extraction

The method used in this study to determine the speciation of radionuclides in the

sediments with the highest uranium content, S3 and S7, was the one that had been

optimized for quantification of actinides in an organic-rich soil (Schultz et al., 1998b).

The only modification was using aqua regia instead of the strong acids (HClO4/HF) in

determining the residual fraction, because the interest in this part of the study is to

examine radionuclide and heavy metal mobility, rather than obtaining the total

concentrations in the residual fraction, and thus allow comparability with the activities

measured by aqua regia leaching.

All reaction steps were performed in duplicate in 50 ml polyethylene centrifuge tubes

with a solid/reagent ratio of 1.0 g/ 15.0 mL. The reagents and conditions for each step

124

are given in Table 4.1. At the beginning of the method, 1.0 g sample was wetted

overnight with water before conducting the extraction steps. To separate the extract

solution following each extraction step, the samples were filtered through 0.22 µm filter

and centrifuged at 4500 rpm/20 minutes. The solid residue was saved for the following

steps and uranium and radium were determined in the solutions. To the solutions,

tracers (232

U for uranium and 225


Th, for radium)

were added before evaporation to dryness using a heating lamp, and uranium and

radium separation was performed.

Table 4.1 Summary of the sequential extraction method applied for radionuclides and

stable elements from Cornwall sediments (sample/reagent ratio is 1.0 g/ 15 .0 mL)

Fraction Extractive reagents Temp. ⁰C Shaking time (h)

Exchangeable 0.4 M MgCl2 R. T. 1

Organic matter 5-6% NaOCl (pH7.5) 96 0.5 x 2

Carbonates 1 M NaOAc in 25% HOAc (pH 4) R. T. 2 x 2

Oxides (Fe/Mn) 0.04 M NH2OH.HCl (pH 2) R. T. 5

Residual HCl/HNO3 (3:1) 96 2

Uranium was separated using extraction chromatography on a UTEVA column, as

described in section 4.2.3.2, while the next section describes the method of radium

separation.

Radiochemical separation of radium was modified from that of (Smith and Mercer,

1970) using 150 mBq 225


Th, as a radiotracer.

Radium was co-precipitated, after adding 50 ml 0.1 M HNO3 to the treated fraction

from the sequential extraction, with PbSO4 by adding consecutively 1.0 mL of

concentrated H2SO4, 2.0 g K2SO4 and 1.0 ml of 0.24 M of Pb(NO3)2. The solid was

125

centrifuged in a 50.0 mL tube at 3000 rpm (about 6200 g) for 10.0 minutes, and then

washed with 20.0 mL of a mixture of 0.2 M H2SO4/0.1 M K2SO4.

The precipitate was dissolved in 5.0 mL of 0.1 M ethylenediaminetetraacetic acid


X8, 100-200 mesh, chloride form, 5 x 0.5 cm) to remove sulphate and washed with 13.0

mL 0.01 EDTA/ NH4OH. To the eluate, 1.0 ml 5.0 M CH3COONH4 was added (pH 4.5)

and the solution was passed through a cation exchange column (Bio-Rad AG50W-X12,

200- 400 mesh, 8.0 x 0.7 cm) at a flow rate of 1.0 mL/minute. The column was

previously conditioned with 15.0 mL 1.5 M CH3COONH4 followed by 15.0 mL 0.25 M

CH3COONH4. Another 50.0 mL 1.5 M CH3COONH4/0.1 M HNO3 was passed through

this column to remove Pb and Ac, while Ba was eluted by washing the column with

40.0 mL 2.5 M HCl. Finally, Ra was eluted with 25.0 mL 6.0 M HNO3, and this

solution was evaporated to dryness using a heating lamp.

The electrolysis cell consists of two glass tubes (Sovril SV 30) joined with a SV 30

plastic joint. A polished stainless steel planchette (cathode) was held between the two

glass tubes by a recessed brass planchette mount supported by the lower electrode. The

cell was sealed with a Teflon ring and checked for leaking. A platinum wire anode,

inserted in a narrow glass tube, was passed through a rubber bung into the electrolyte

solution to complete the electric circuit.

For radium electroplating, the Ra fraction was re-dissolved in organic electrolyte

solution (1.0 mL 0.1 M HNO3 in 9.0 mL ethanol) and electroplated on to a stainless

steel planchette at 120 mA for 30 minutes. One minute before the end, 1.0 ml of

ammonia (s.g. 0.88) was added and the power was turned off. The solution was

discarded and the planchette was dried on a hotplate at 200 °C for 5 minutes.

126

4.2.4.3 Stable element analysis

The leachates from sequential extraction of the sediments with the highest uranium

content, S3 and S7, and the 20 water samples were analysed using inductively coupled

plasma atomic emission (ICP-MS) and inductively coupled plasma mass spectrometry

(ICP-MS), the latter where the concentrations were low enough to require it. For

analysis of the sediment samples, 1 ml of each fraction obtained from the sequential

extraction was made up to 10 mL with 2 % nitric acid. The samples were run using a

series of solutions prepared from certified standard solutions (1000 ppm, Sigma

Aldrich, UK) for each analyte. Water samples for cation measurements were filtered

and acidified in the field with nitric acid to a pH < 2, then analysed directly.

4.2.5 Sediment characterisation

4.2.5.1 Heavy liquid separation

The heavy liquid separation technique was used to separate minerals in the sediments

with the highest uranium content, S3 and S7, based on density. The sample was placed

into a 50.0 mL centrifuge tube and a heavy liquid for density separation, LST Fastfloat,

which consists of sodium heteropolytungstates dissolved in water to give a density 2.80

± 0.02 g/mL, was poured to half-fill the tube. The tube was hand-shaken to mix the

grains with the heavy liquid then more LST was added until the tube was almost full,

and left overnight to allow the minerals to separate. Once the minerals had separated,

the lower end of the centrifuge tube was immersed into a small container of liquid

nitrogen until the bottom 1 cm of liquid was frozen. The unfrozen solution was decanted

and filtered under gravity. Deionised water was used carefully to rinse out any minerals

remaining in the tube, while avoiding melting the frozen layer. The bottom layer was

then allowed to melt, and filtered under gravity, rinsing with deionised water. The

filtered samples were rinsed 4-5 times with deionised water to ensure removal of LST,

and the filter papers were placed inside an oven to dry at 100 °C.

127

4.2.5.2 Scanning electron microscopy analysis

Two sediment samples with the highest radioactivity, S3 and S7, were selected for

characterisation by the JEOL JSM-6400 SEM. Three subsamples (total sample, light

minerals and heavy minerals) were prepared. Each dry sample was embedded on a glass

slide using epoxy resin, and polished to provide a homogeneous surface for analysis.

The samples were carbon-coated so the samples were conductive, to prevent charging of

the surface and to promote emission of secondary electrons. At the beginning of the

analysis, backscattered images were obtained to localise heavy elements. This was

followed by obtaining secondary electron images from the near surface of the most

interesting spots using a voltage of 15-20 kV. In addition, the EDX Princeton Gamma

Tech EDS system was used to perform semi-quantitative elemental analysis.

4.2.5.3 Electron microprobe analysis

The same two sediment samples (S3 and S7) that were characterised by SEM were

further analysed using a CAMECA SX100 electron microprobe analyser (EMPA).

Firstly, a backscattered image of a 100 x 100 µm area of the sample was obtained, to

locate higher atomic number elements. This was followed by selecting a single grain

(typically 50 x 50 µm) for elemental mapping. In addition, the elemental composition

(expressed as the oxides) of 21 major and trace elements was determined by energy-

dispersive spectroscopy at selected spots. During analysis, the acceleration voltage was

15 kV and the beam current of the probe was 20 nA, and several standards were used

for calibration.

The instrument is equipped with five wavelength detectors and it was also possible to

use these to obtain elemental maps using wavelength-dispersive spectroscopy. The 10

elements selected were divided into two groups. The first group included U, Ca, Mg,

Mn and Fe, while K, Cu, As, Sn and Pb were in the second group.

128


4.3.1 Physico-chemical properties of stream waters

Sampling location coordinates, water physico-chemical data and anion concentrations of

the 20 water samples collected from the River Fal and tributary streams are presented in

Table 4.2. The water pH shows a slightly acidic range of 5.85-7.00, as expected in

rivers draining on granitic bedrocks (Carvalho et al., 2007). The other water quality

parameters, such as conductivity, total dissolved solids, chloride, nitrate and sulphate

for the samples, are within the fresh water guideline values provided by the World

Health Organisation (WHO, 1998).

Table 4.4 presents selected major and trace elements concentrations in the water

samples. For all elements, the data are within the values recommended by the WHO.

For uranium, the average concentration of the samples collected from the River Fal is

0.2 µg/L, which is in good agreement with the average global concentration of river

water (Palmer and Edmond, 1993). However, at sampling points S1, S2, S3 and S6, the

uranium values are 2-4 times higher than the background. S7 displays the highest

concentration of uranium, in addition to relatively high copper and arsenic

concentrations, maybe reflecting downstream transport of particulate material.

4.3.2 Physico-chemical properties of sediments

The bulk minerals identified by X-ray diffraction (XRD) and the organic matter (OM)

content calculated from the loss of ignition are outlined in Table 4.5. The OM content of

the River Fal sediment is much lower (1.0%) than that of the sediments collected from

streams running towards the river. In some streams, the OM content of the sediments is

as high as 37% (S3), 25% (S2) and 21% (S7). Organic-rich sediments are also observed

in some streams emerging from adits (e.g. 29% in S12 and 20% in S13) along the river

course.

129

XRD is only useful in characterising crystalline minerals, so minerals with disordered

structures will not give well defined patterns. The mineralogical composition of the bulk

samples reveals that quartz is the main component of all sediments, as expected in

stream sediments. In addition muscovite, kaolinite, rutile and schorl are identified in all

samples. It is important to note that the lowest percentage of quartz was found in the

sample with the highest uranium content (S7), and the highest proportion of

phyllosilicate minerals (muscovite and kaolinite). Minor dolomite and chlorite were also

identified, addition to jarosite (found in one sample coming from an adit). Overall

though, the results from XRD suggest no substantive difference in bulk sediment

composition, whether for those collected from the main river or from the side streams.

4.3.3 Radiochemical characterisation of sediments

The activity concentrations and activity ratios of the natural radionuclides 238

U, 234

U and

226Ra in the sediment from the River Fal and streams around South Terras mine are

presented in Table 4.6.

The sample located about 100 m upstream (S20) was selected to provide a background

level for the sediments in the River Fal. This point shows negligible influence from the

mine on the uranium activity (~72 Bq kg-1

), while the concentration of radium is 51 Bq

kg-1

. Moreover, the ratios of 234

U/238

U and 226

Ra/238

U in S20 are close to those observed

far from uranium mines (Lozano et al., 2002a). The highest concentrations of the

radionuclides are found in the sediments collected downstream, about 135 m south of

the mine building (S7). The values are up to 4350, 4265 and 1765 Bq kg-1

for 238

U, 234

U

and 226

Ra, respectively. The ratio between the two uranium isotopes is around one,

indicating equilibrium.

At a distance of about 425 m downstream, the concentrations of the radionuclides vary

in the small streams draining towards the River Fal. For samples S1 and S2, the activity

130

concentrations of 238

U are 230 and 290 Bq kg-1

. These samples were depleted slightly in

the daughter, 234

U, compared with the parent, 238

U. In the same area, sample S3 gave a

238U content of 1820 Bq kg

-1 and a similar depletion in

234U to S1 and S2. In addition,

the 226

Ra shows activity concentration higher than the background in the sediments

from these streams, with a value of 940 Bq kg-1

at S3.

For the remainder of the samples, with the exception of the 226

Ra level in S6 and S12,

the sediments in the River Fal seem to display local background levels (range 42 to 115

Bq kg-1

; mean value 64 Bq kg-1

). Since sediments can indicate the source of

contamination in an area, the results from the River Fal suggest that the South Terras

mine has no significant radiological impact on the nearby water courses beyond a

distance greater than 0.5 km from the mine buildings.

4.3.4 Sequential chemical extraction results

Elevated activity concentrations of 238

U, 234

U and 226

Ra were observed in the two

samples (S3 and S7) collected close to the mine building. In order to identify the

association of the radionuclides with the different geochemical fractions, sequential

chemical extraction (SCE) was performed.

4.3.4.1 Fractionation in S3:

The sequential extraction results of uranium (Fig. 4.2) showed that about 43 % of the

total uranium was associated with Fraction 2, interpreted as the organic fraction, and

around 55 % of the uranium was extracted in fraction 3, interpreted as the carbonate

phase. Carbonate species, most commonly CaCO3, represent an effective uranium sink

(Kipp et al., 2009; Rachkova et al., 2010). These two fractions represent about 95% of

the uranium released in all fractions, except resistant, suggesting strong retention of

uranium by adsorption to organic species and/or incorporation into carbonates.

131

For radium, as can be seen in Fig. 4.3, about 70% was found to be associated with

Fraction 2, the organic fraction, while about 15% was released in Fraction 3, the

carbonate fraction, and only 10% was released in Fraction 1, the exchangeable fraction.

This may be linked to the higher organic matter (37%) in this sediment (Greeman et al.,

1999).

4.3.4.2 Fractionation in S7:

Around 60% of the total uranium was extracted in Fraction 3, the carbonate fraction

(Fig. 4.2), the same as S3, while 25% of the uranium was extracted from Fraction 5, the

resistant fraction and only about 15% from Fraction 2, the organic fraction.

This is again consistent with the proposed association between uranium and carbonate

phases. However, compared to sample S3, a higher percentage of uranium was present

in the residual fraction, Fraction 5, suggesting the presence of primary U-minerals in

S7.

For radium, 60% was found to be associated with Fraction 2, the organic fraction, and

20% was released in Fraction 5, the residual fraction (Fig. 4.3). In contrast to S3, part of

the radium appears to be associated with a resistant phase, rather than primarily

controlled by adsorption. This would be reasonable, given the presence of uranium in

the residual fraction and therefore the potential for in situ generation of radium.

4.3.5 Radionuclide and stable element fractionation

Comparing the geochemical distribution of the stable elements with the radionuclides in

S3 and S7 could help understanding of fractionation within the operationally-defined

phases. Accordingly, geochemical fractionation of selected stable elements was

undertaken in order to explore the most likely geochemical host phases for uranium and

radium in Cornwall sediments.

132

Calcium was associated mainly with the Fractions 1 and 2 (exchangeable and organic)

fractions in both samples (Fig. 4.4), with the majority (~ 80 %) of the calcium in S7

bound to Fraction 2, compared with only 60% in S3.

Manganese (Fig. 4.5) was also associated primarily with Fractions 1 and 2, the

exchangeable and organic matter fractions, in S3, with nearly the same percentage in

each fraction. However, in S7 it was found in all fractions, with a relatively higher

percentage (~ 40%) in Fraction 2 (organic).

From Fig. 4.7, arsenic was found to associate with Fractions 2 and 3 of S7, with ~ 60%

in carbonate and ~ 37% in organic. In S3, it was more distributed towards the acid-

resistant fractions, with about 30% in Fraction 5, the resistant component.

Barium fractionation (Fig. 4.9) is very similar to that of Ra. In S3, it was attached

predominantly to Fraction 2; however, for S7 it was distributed in all fractions, with the

exception of Fraction 4 (Fe/Mn oxides), with a relatively higher amount (~ 40%) in

Fraction 5, the resistant fraction.

Fractionation of titanium and iron are illustrated in Figs. 4.6 and 4.8. In both samples,

the majority of titanium (~ 90%) was released from Fraction 5, the fraction leached by

strong acids. This is consistent with the classification of Ti as a refractory element

(Schultz et al., 1998a), which, therefore, is expected to be bound to the resistant

fraction. In both sediments, iron, same as titanium, was also found mainly in this

fraction, with about 75%.

Comparing the geochemical fractionation of stable elements with that of uranium and

radium showed that, in both S3 and S7, the uranium distribution was quite different

from that of any stable elements. In both sediments, the fractionation profile of barium

133

showed strong similarities to that of radium. Again, this is expected since barium is a

close chemical analogue of radium.

4.3.6 Uranium isotopic ratios in sequential extraction fractions

The activity ratio between 238

U and 234

U isotopes was used to obtain information about

the mobility of uranium. The uranium isotopic activity ratio (234

U/238

U) in the sequential

extraction fractions of samples S3 and S7, excluding the exchangeable fraction, is

shown in Figs. 4.10 and 4.11.

For S3, the results indicated that the 234

U/238

U activity ratio in the uranium in Fraction 2,

the organic fraction, was close to unity; however, in the remaining fractions, uranium

isotopic ratios were lower than unity. As mentioned above, S3 has the highest OM

content of the Cornwall sediments (~ 40%) and OM plays an important role in retaining

radionuclides. A possible reason for this higher 234

U/238

U in the organic fraction

compared with the other fractions is that the organic fraction, which will include organic

coatings on other particles, is more likely to be in contact with water, and therefore has

a great opportunity to adsorb uranium from water, with an activity ratio greater than

unity (Vargas et al., 1997).

For S7, Fractions 3 and 4, the carbonates and Fe/Mn oxides, revealed 234

U/238

U activity

ratios close to unity, while the resistant fraction showed the uranium isotopes were in

equilibrium. However, the 234

U/238

U activity ratio in Fraction 2, the organic fraction,

was ca 0.8, indicating substantial disequilibrium. As discussed earlier, disequilibrium in

sediments is interpreted as relating to the preferential leaching of 234

U from the mineral

grain, compared with the parent 238

U. Consequently, water is expected to exhibit

234U/

238U > 1, while sediment is expected to show

234U/

238U < 1 (Riotte and Chabaux,

1999; Vigier et al., 2005; Vigier et al., 2006). In S7, which has the highest uranium

content of the Cornish sediments, the equilibrium 234

U/238

U ratio may reflect the

134

presence of significant primary U minerals (hence the relatively high proportion of

uranium in Fraction 5, the resistant component), and this material could contribute

uranium to Fractions 3 and 4, giving the same ratio. By contrast, the much lower ratio in

Fraction 2 suggests a more complex process, for example, maybe, derivation of the

uranium in this fraction from another source.

4.3.7 Characterisation of sediments using spectroscopic methods

4.3.7.1 Scanning electron microscopy

In the scanning electron microscopy (SEM) analysis, the bright areas in a backscattered

electron image (BSE) indicate the presence of high atomic number elements, which

could be uranium or other heavy elements, such as Ti. However, the electron-dispersive

spectroscopy (EDS) spectrum provides a semi-quantitative analysis of the area of

interest which allows identification of the element in the area. The identification of U-

bearing particles in S3, without heavy liquid separation, is illustrated in Figures 4.12

and 4.13. The BSE image identifies a bright spot, as demonstrated in Fig. 4.12,

indicating the presence of heavy element-bearing particles. The chemical composition

of this bright area from EDS analysis suggests the existence of trace amounts of

uranium and other metals (Al, Si, P, K and Ca). As the scale of the area being imaged

becomes smaller (~ 20 µm), focusing more on the bright area (Fig. 4.13), the intensity

of the uranium signal increases and the those of the associated elements decrease. By

applying secondary electron (SE) analysis in combination with EDS on this small scale,

it is possible to localise an individual U-bearing particle. Since the SE image is derived

from low energy (< 50 keV) electrons with limited penetrating ability, the uranium is

expected to be very close to the surface of the particle, possibly in a coating layer.

The SEM/EDS results of the heavy minerals of S7, isolated using heavy liquid

separation, revealed the presence of U-bearing particles, as presented in Figures 4.14

135

and 4.15. The EDS spectrum (Figure 4.14) identified an association of uranium with P,

Ca and Th in one particle. In another particle (Figure 4.15), the uranium was found to

associate with clay minerals (Al and Si). This may support the sequential extraction

results, where a considerable amount (25%) of uranium in S7 was released from the

residual fraction, while, for S3, only 1% of the uranium was released from same

fractions.

4.3.7.2 Electron microprobe analyser

In an attempt to locate uranium hot spots in the minerals separated by heavy liquid from

S3 and S7, an electron microprobe analyser was used for further characterisation. As in

SEM analysis, BSE imaging was used to identify areas with higher atomic number

elements relative to the adjacent areas. Wavelength-dispersive spectrometry (WDS) was

used to create X-ray maps of 10 elements (K, Mg, Ca, Mn, Fe, Cu, As, Sn, Pb and U) in

order to obtain the chemical distribution within the grain of interest. The results is

presented in Figures 4.16.

For S3, electron microprobe analyser (EMPA) characterisation of the bulk minerals

could not identify a grain containing uranium. However, for S7, a grain from the heavy

minerals separated by the heavy liquid was identified as containing trace amounts of

uranium-bearing minerals (Fig. 4.16). The WDS images suggested that the grain

includes a higher amount of K, Ca and Fe relative to U, As and Mn. The percentage of

uranium oxide (wt.% UO2) in this grain was about 1%.

4.4 Conclusions

Radioactivity around the former uranium mining site at South Terras is generally close

to local background levels, with no substantial effect of the radionuclides on the River

Fal, and enhanced concentrations of radionuclides only found in the immediate area of

the mine. The elevated activity at distances less than 0.5 km could be related to the

136

migration of particles enriched in uranium from the mine building due to the weathering

effect.

Sequential chemical extraction, applied to the sediments with the highest radionuclide

concentrations, revealed different geochemical fractionation of uranium and radium.

The uranium in the sediment with the highest organic matter was more closely

associated with relatively labile fractions, particularly the organic and the carbonate

fractions. Furthermore, the sample with the highest uranium revealed that although the

carbonate bound more of the uranium, a significant portion was held in the resistant

fraction. Radium in both sediments was held primarily in the organic fraction but, in S7,

the sample with the highest radioactivity, significant radium was also held in the

resistant fraction.

There was no clear association of the radionuclides and the stable elements in individual

fractions, although there were indications of an association with calcium, manganese

and arsenic in the organic and the carbonate fractions..

The activity ratios of the uranium isotopes may suggest that exchange between

sediment/water could explain the findings so further study, including measurements of

the uranium isotopic ratios in water samples, is recommended.

Using SEM, uranium-bearing particles have been localised in the bulk minerals and the

heavy minerals from the sediments enriched in uranium. EMPA results produced an X-

ray map of the uranium, with associated stable elements, in a single grain obtained from

the sample with the highest uranium content.

Acknowledgements

The authors greatly appreciate financial support from the Islamic Development Bank

(IDB), Jeddah, Saudi Arabia. The authors thank Mr Paul Lythgoe for ICP-OES/ICP-MS

137

measurements, Mr Alastair Bewsher for IC analysis, Mrs Cath Davies for total

dissolution of samples, Dr John Waters for SEM and Dr John Charnock for EMPA

analysis, all from School of Earth, Atmospheric and Environmental Sciences,

University of Manchester.

138

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Sediment and radionuclide transport at the former Nabarlek uranium mine, Northern

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U/238

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142

Table 4.2 Physiochemical properties, anions of water samples collected from the River

Fal and side streams in Cornwall and the coordinates of the sampling points

ID pH EC

µS/cm

TDS

ppm

Cl-

mg/L

NO3-

mg/L

SO42-

mg/L

Latitude

(N)

Longitude

(W)

S1 6.00 264 177 34.7 62.1 3.1 50⁰ 19.861' 4⁰ 54.248'

S2 6.20 270 181 35.1 50.3 3.2 50⁰ 19.863' 4⁰ 54.257'

S3 6.15 273 183 35.1 62.9 3.1 50⁰ 19.856' 4⁰ 54.248'

S4 6.80 222 149 23.4 10.7 21.9 50⁰ 19.809' 4⁰ 54.274'

S5 6.80 223 149 23.4 10.2 22.2 50⁰ 19.810' 4⁰ 54.272'

S6 6.25 271 182 35.0 62.1 3.2 50⁰ 19.809' 4⁰ 54.275'

S7 6.60 407 273 41.1 76.5 11.3 50⁰ 20.014' 4⁰ 54.330'

S8 6.15 260 174 29.5 31.1 9.6 50⁰ 20.059' 4⁰ 54.344'

S9 6.80 237 159 24.5 11 23.1 50⁰ 20.103' 4⁰ 54.321'

S10 6.90 233 156 24.5 11.5 22.9 50⁰ 19.559' 4⁰ 54.259'

S11 7.00 235 157 24.7 11.5 23.2 50⁰ 19.559' 4⁰ 54.260'

S12 5.85 231 155 31.2 46.4 1.7 50⁰ 19.558' 4⁰ 54.260'

S13 5.95 227 152 29.3 50.8 1.5 50⁰ 19.547' 4⁰ 54.233'

S14 6.95 237 159 24.7 11.7 24.1 50⁰ 19.550' 4⁰ 54.233'

S15 7.00 235 157 24.5 11.5 24.4 50⁰ 19.014' 4⁰ 54.330'

S16 7.00 238 159 24.8 11.7 25.6 50⁰ 19.359' 4⁰ 54.019'

S17 6.35 218 146 31.8 1.0 23.7 50⁰ 19.371' 4⁰ 54.010'

S18 6.85 238 159 24.9 11.8 26.5 50⁰ 19.422' 4⁰ 54.097'

S19 6.90 239 160 24.5 11.3 27.4 50⁰ 19.701' 4⁰ 54.286'

S20 6.95 240 161 23.9 9.8 28.8 50⁰ 20.138' 4⁰ 54.352'

143


Ra and 238

U in

IAEA-314 stream sediment reference material

226

Ra Bq.kg-1

238

U mg.kg-1

Measured 774 ± 24 57.8

Recommended 732 56.8

95% Confidence interval 678 – 787 52.9 – 60.7

144

Table 4.4 Concentrations of cations in mg/L (g/L for Cu, As, Pb and U) in the filtered water samples (<0.22 m) collected from the River

Fal and side streams in Cornwall

ND = Not detected

ID Na K Mg Ca Ba Mn Fe Zn Cu As Pb U

W1 10.98 1.97 14.61 6.37 0.01 0.02 0 0.107 3.70 3.46 0.15 0.74

W2 11.08 2.37 14.42 6.66 0.01 0.12 0.034 0.096 2.60 3.70 0.07 0.46

W3 10.84 2.01 14.76 6.43 0.01 0.02 0.002 0.106 3.73 3.51 0.13 0.84

W4 12.8 3.7 5.51 9.36 0.01 0.07 0.391 0.003 1.81 2.42 0.04 0.21

W5 13.03 3.83 5.615 9.40 0.01 0.07 0.394 0.003 1.84 2.30 0.04 0.21

W6 10.8 1.99 14.66 6.39 0.01 0.02 0.002 0.102 1.86 2.88 ND 0.38

W7 16.28 5.36 13.68 26.26 0.03 0.013 0.006 0.07 10.05 25.35 ND 3.69

W8 9.879 5.63 9.827 13.04 0.02 0.17 0.553 0.022 2.02 3.22 0.20 0.03

W9 13.64 3.98 5.624 9.33 0.01 0.08 0.332 0.004 1.66 2.16 0.02 0.16

W10 13.51 3.89 5.783 9.38 0.01 0.07 0.326 0.004 1.72 2.25 0.02 0.20

W11 13.73 3.94 5.907 9.59 0.01 0.07 0.322 0.006 2.36 2.16 0.02 0.19

W12 9.891 0.74 12.64 5.21 0.003 0.11 0.003 0.268 0.99 0.17 0.06 <0.001

W13 9.48 0.55 11.73 5.44 0.004 0.01 0.002 0.080 1.01 <0.05 0.52 0.002

W14 13.52 3.85 5.866 9.49 0.01 0.07 0.352 0.005 2.00 2.19 0.06 0.19

W15 13.42 3.83 5.846 9.40 0.01 0.07 0.352 0.006 1.98 2.31 0.06 0.19

W16 13.57 3.90 5.843 9.45 0.01 0.07 0.353 0.005 3.92 2.09 0.50 0.20

W17 11.5 2.84 9.775 11.52 0.01 0.01 0.036 0.735 5.10 1.42 12.78 0.03

W18 13.7 4.0 5.883 9.55 0.01 0.07 0.342 0.007 3.23 2.12 0.07 0.20

W19 13.56 4.07 5.847 9.56 0.01 0.07 0.345 0.005 3.41 2.26 0.03 0.19

W20 14.15 4.05 5.734 10.76 0.011 0.08 0.386 0.005 3.49 2.29 0.08 0.17

145

Table 4.5 Mineralogical composition from XRD and loss on ignition of sediments

collected from the River Fal and side streams in Cornwall

ID L.O.I. Quartz% a Total phy.% TiO2% Dolomite% Chlorite% Jarosite%

S1 5.2 88 6 5 - - -

S2 22.3 77 18 1 - 4 -

S3 37.0 77 11 7 - 2 -

S4 1.0 84 12 2 - 2 -

S5 1.0 50 11 6 1 1 -

S6 5.8 76 15 8 1 - -

S7 21.0 66 20 9 2 3 -

S8 24.8 79 7 6 1 7 -

S9 2.1 80 10 7 1 2 -

S10 0.7 79 11 6 1 2 -

S11 1.0 73 10 7 1 2 -

S12 28.6 82 7 6 1 2 -

S13 20.0 80 12 6 - 2 -

S14 1.5 80 12 6 2 - -

S15 0.7 78 11 9 1 1 -

S16 0.8 73 16 9 2 - -

S17 5.8 77 5 8 - 3 5

S18 0.8 74 15 9 2 - -

S19 1.1 79 10 8 1 1 -

S20 1.0 87 10 5 1 2 -

a Total phy. Represents total phyllosilicates (muscovite and kaolinite)

146

Table 4.6 U-isotopes and Ra activity concentrations (Bq kg-1

dry weight) and isotopic

ratios in 20 sediment samples collected from locations around the River Fal and side

streams in Cornwall (± 1σ counting statistics uncertainties)

ID 238

U 234

U 226

Ra 234

U/238

U 226

Ra/238

U

S1 230 ± 12 181 ± 10 293 ± 15 0.78 ± 0.06 1.27 ± 0.09

S2 290 ± 14 260 ± 13 424 ± 23 0.89 ± 0.06 1.45 ± 0.10

S3 1820 ± 36 1388 ± 32 940 ± 53 0.76 ± 0.02 0.52 ± 0.03

S4 43 ± 5 44 ± 5 55 ± 6 1.02 ± 0.18 1.27 ± 0.20

S5 49 ± 6 55 ± 6 68 ± 8 1.12 ± 0.19 1.38 ± 0.23

S6 160 ± 10 127 ± 9 220 ± 12 0.80 ± 0.08 1.40 ± 0.12

S7 4350 ± 53 4265 ± 52 1765 ± 48 0.98 ± 0.02 0.41 ± 0.01

S8 95 ± 8 104 ± 8 116 ± 11 1.10 ± 0.13 1.22 ± 0.15

S9 43 ± 5 39 ± 5 75 ± 7 0.90 ± 0.16 1.75 ± 0.26

S10 51 ± 5 57 ± 5 61 ± 5 1.12 ± 0.14 1.21 ± 0.15

S11 40 ± 6 40 ± 6 61 ±5 0.98 ± 0.19 1.51 ±0.24

S12 39 ± 5 35 ± 5 194 ± 16 0.88 ± 0.17 4.92 ±0.75

S13 44 ± 5 54 ± 6 72 ± 7 1.22 ± 0.20 1.64 ± 0.26

S14 50 ± 6 49 ± 6 67 ± 6 0.99 ± 0.16 1.35 ±0.19

S15 49 ± 6 52 ± 6 60 ± 5 1.06 ± 0.18 1.23 ±0.18

S16 44 ± 6 42 ± 6 53 ± 4 0.95 ± 0.18 1.21 ± 0.19

S17 13 ± 3 16 ± 4 42 ± 4 1.22 ± 0.39 3.12 ± 0.80

S18 43 ± 6 42 ± 6 62 ± 5 0.98 ± 0.18 1.45 ±0.21

S19 41 ± 5 40 ± 5 53 ± 4 0.98 ± 0.19 1.32 ±0.22

S20 73 ± 7 66 ± 6 51 ± 4 0.91 ± 0.12 0.71 ± 0.09

147

Figure 4.1 Cornwall map showing the sampling points along the river Fal

148

Figure 4.2 Extraction profile of uranium as a percentage of the sum of five fractions in

S3 and S7

Figure 4.3 Extraction profile of radium as a percentage of the sum of five fractions in

S3 and S7

0.00

10.00

20.00

30.00

40.00

50.00

60.00

70.00

80.00

90.00

100.00

Exchangeable Organic Carbonate Fe/Mn oxides Resistant

Pe

rce

nta

ge

Sequential extraction fractions

S3

S7

0.00

10.00

20.00

30.00

40.00

50.00

60.00

70.00

80.00

90.00

100.00

Exchangeable Organic Carbonates Fe/Mn oxides Resistant

Per

cen

tage


S3

S7

149

Figure 4.4 Extraction profile of calcium as a percentage of the sum of five fractions in

S3 and S7

Figure 4.5 Extraction profile of manganese as a percentage of the sum of five fractions

in S3 and S7

0.00

10.00

20.00

30.00

40.00

50.00

60.00

70.00

80.00

90.00

100.00


Pe

rce

nta

ge


Ca S3

Ca S7

0.00

10.00

20.00

30.00

40.00

50.00

60.00

70.00

80.00

90.00

100.00


Per

cen

tage


S3

S7

150

Figure 4.6 Extraction profile of iron as a percentage of the sum of five fractions in S3

and S7

Figure 4.7 Extraction profile of arsenic as a percentage of the sum of five fractions in

S3 and S7

0.00

10.00

20.00

30.00

40.00

50.00

60.00

70.00

80.00

90.00

100.00


Pe

rce

nta

ge


S3

S7

0.00

10.00

20.00

30.00

40.00

50.00

60.00

70.00

80.00

90.00

100.00


Per

cen

tage


S3

S7

151

Figure 4.8 Extraction profile of titanium as a percentage of the sum of five fractions in

S3 and S7

Figure 4.9 Extraction profile of barium as a percentage of the sum of five fractions in

S3 and S7

0

10

20

30

40

50

60

70

80

90

100


Pe

rce

nta

ge


S3

S7

0.00

10.00

20.00

30.00

40.00

50.00

60.00

70.00

80.00

90.00

100.00


Per

cen

tage

Sequential extraction fraction

S3

S7

152

Figure 4.10 234

U/238

U activity ratios in the sequential extraction fractions of S3

0.00

0.20

0.40

0.60

0.80

1.00

1.20

Organic Carbonate Fe/Mn oxides Resistant

23

4U

/24

8U

Act

ivit

y r

atio


153

Figure 4.11 234

U/238

U activity ratios in the sequential extraction fractions of S7

0.00

0.20

0.40

0.60

0.80

1.00

1.20

Organic Carbonate Fe/Mn oxides Resistant

23

4U

/24

8U

Act

ivit

y r

atio


154

Figure 4.12 Scanning electron microscope (SEM) results showing backscattered

electron (BSE) image (A) and electron-dispersive spectroscopy (EXD) spectrum (B) of

the bulk minerals of S3. The bright area is an indication of a presence of high atomic

number element.

A

B

155

Figure 4.13 Scanning electron microscope (SEM) results showing secondary electron

(SE) image (A) and electron-dispersive spectroscopy (EXD) spectrum (B) of the bulk

minerals of S3.

A

B

156


(SE) image (A) and electron-dispersive spectroscopy (EXD) spectrum (B) of a single

grain isolated from the heavy minerals separated by heavy liquid from the richest U-

sample (S7). U association with P, Th and Ca (from the EDX analysis) has been

identified.

A

B

157


(SE) image (A) and electron-dispersive spectroscopy (EXD) spectrum (B) of the heavy

minerals separated by heavy liquid fractionation of S7. U associates with the

aluminosilicates has been identifird .

A

B

158

Figure 4.16 Backscattered electron (BSE) image (top) and X-ray maps of elements

(Mn, K, Fe, As, Ca and U) from electron microprobe analysis (EMPA) of the heavy

minerals separated by heavy liquid of S7

159

Chapter Five

Conclusions and Recommendations

An analytical method for 226

Ra separation using 225

Ra as a radiotracer, in equilibrium

with 229

Th, by alpha spectrometry has been developed. The obvious advantage of the

method is that it is direct, simple and accurate, since a radium isotope was used as a

yield determinant. In comparison with the methods using 133

Ba as a spike, assuming the

similarity in chemistry between Ra/Ba, the method operated in this project presents

more accurate measurement. Moreover, the method enables determination of all four

naturally occurring radium isotopes in the same prepared alpha source, after a suitable

ingrowth time for 228

Ra.

The aim of the second part of the project was to understand chemical weathering and

physical erosion impact on the natural U-series in the valley of the River Noe, Edale.

The radionuclide concentrations in stream sediments from the valley, applying total

dissolution and aqua regia leaching methods, showed considerable variation in 238

U,

234U,

230Th and

226Ra content. The α-decay process in addition to accumulation/leaching

of the parent materials, organic matter and secondary minerals within the sediments

appeared to derive the radionuclides’ fractionation. The isotopic ratios suggested

complex U-series transport, and this was reflected in a plot of 234

U/ 238

U against 230

Th/

238U. Organic matter and secondary minerals affect the distribution of U-series

radionuclides by adsorbing uranium and radium onto mineral surfaces and/or

complexation with thorium. Cluster analysis provides insight into radionuclide

behaviour and suggests a relationship between the landslips in the River Noe Valley and

the stream sediment isotope concentrations. It is possible that uranium-containing

material has been exposed by the slips, and that 226

Ra and 234

U are being released into

the runoff water, and then becoming sorbed onto the stream sediments.

160

Radioactivity around the former uranium mining site at South Terras is generally close

to local background levels, with no substantial effect of the radionuclides on the River

Fal, and enhanced concentrations of radionuclides only found in the immediate area of

the mine. The elevated activity at distances less than 0.5 km could be related to the

migration of particles enriched in uranium from the mine locality.

Sequential chemical extraction results showed different geochemical distribution of

uranium and radium in the sediments. Uranium was more adsorbed to Fraction 2

(organic) and Fraction 3 (carbonate) in the organic-rich sediment, while Fraction 3

(carbonate) and Fraction 5 (resistant) bound more of uranium in sediment with the

highest U-content. Both sediments attached radium to Fraction 2 (organic), although in

the sample with the elevated radioactivity, significant radium was also held in Fraction

5 (resistant). There were similar geochemical distributions between uranium, calcium,

manganese and arsenic in Fraction 2 (organic) and Fraction 3 (carbonate). 234

U/238

U

activity ratios in the sequential extraction fractions suggested different degrees of

equilibration between sediment and water.

Scanning electron microscopy analysis identified uranium-bearing particles in bulk

minerals and heavy minerals, separated by heavy liquid, from the uranium-rich

sediments. Electron microprobe analysis localised uranium in the uranium-rich sample.

161

Future work

In the Edale study, the results suggest complex U-series disequilibrium in sediments, so

this could be further investigated proposing the following:

Sequential chemical extraction of sediments close to the landslips, to

determine phase associations of the radionuclides.

Sampling of the landslip areas themselves.

Physico-chemical analysis of water from the study area, particularly

groundwater and streams adjacent to the landslips.

For Cornwall, the results reveal a very limited effect of South Terras mine on the

surrounding environment. In particular, there is no substantial radiological impact from

the mine in the sediment beyond 500 m distance, nor is there contamination in the River

Fal water. Future investigation may include:

Autoradiography in order to identify sediments rich in radioactive particulates.

Measurements of uranium isotopic ratios in water samples may be by multi-

collector inductively coupled plasma spectroscopy (MC-ICP-MS), to explore the

accumulation/leaching of uranium and water/sediment exchange processes.

162

Appendix

Methods and experimental techniques

A1 Areas of the study

A1.1 Edale sampling

Twenty-five surface sediments were collected by hand from accessible points of the

streams draining on both sides along the River Noe, in the Edale Valley of the Peak

District, during two field trips (10 and 17 December 2010) and saved in Kraft® paper

envelopes. In the laboratory, the sediments were wet-sieved through 2 mm mesh and

left to air dry on open trays for several days. The dry sediments were gently

disaggregated using a mortar and pestle, and stored in plastic bottles.

A1.2 Cornwall sampling

Twenty locations along an approximately 2 km stretch of the valley of the River Fal,

running south from the South Terras mine site, were sampled for water and surface

sediments (17-20 May 2011).

The water samples were collected in polyethylene bottles. As soon as possible after

collection (always within 12 hours), each water sample was divided into three

subsamples: unacidified, unfiltered (for physicochemical analysis, such as pH and

electrical conductivity); acidified, filtered (for elemental analysis); and unacidified,

filtered (for anion and total dissolved carbon measurement). The filtration was

conducted using 0.22 µm cellulose acetate filters and the acidification was done using

nitric acid (1 ml concentrated HNO3 per 100 ml of water). In the field, the samples’ pH

was measured using a pH-meter (SevenEasy, Mettler-Toledo GmbH).

The sediment samples were saved in Kraft® paper envelopes. In the laboratory, the

sediments were wet-sieved through 2 mm mesh and left to air dry on open trays for

163

several days. The dry sediments were gently disaggregated using a mortar and pestle

and stored in plastic bottles.

164

Table A1 Sample locations from the River Noe in the Edale Valley, the Peak District

Sample Coordinates

E1 53⁰ 21.558' N; 1⁰ 50.107' W

E2 53⁰ 21.549' N; 1⁰ 50.042' W

E3 53⁰ 21.526' N; 1⁰ 49.722' W

E4 53⁰ 21.518' N; 1⁰ 49.561' W

E5 53⁰ 21.513' N; 1⁰ 49.567' W

E6 53⁰ 21.518' N; 1⁰ 49.276' W

E7 53⁰ 21.508' N; 1⁰ 49.187' W

E8 53⁰ 21.475' N; 1⁰ 48.903' W

E9 53⁰ 21.952' N; 1⁰ 49.350' W

E10 53⁰ 21.895' N; 1⁰ 49.458' W

E11 53⁰ 21.730' N; 1⁰ 49.711' W

E12 53⁰ 21.696' N; 1⁰ 49.840' W

E13 53⁰ 22.108' N; 1⁰ 48.893' W

E14 53⁰ 22.221' N; 1⁰ 48.488' W

E15 53⁰ 22.178' N; 1⁰ 48.350' W

E16 53⁰ 22.239' N; 1⁰ 47.947' W

E17 53⁰ 22.211' N; 1⁰ 48.009' W

E18 53⁰ 21.563' N; 1⁰ 50.471' W

E19 53⁰ 21.648' N; 1⁰ 50.660' W

E20 53⁰ 21.882' N; 1 50.842' W

E21 53⁰ 21.875' N; 1⁰ 50.912' W

E22 53⁰ 21.769' N; 1⁰ 48.406' W

E23 53⁰ 21.735' N; 1 48.678' W

E24 53⁰ 21.783' N; 1⁰ 48.303' W

E25 53⁰ 21.798' N; 1⁰ 48.237' W

165

Table A2 Sample locations from the valley of the River Fal, Cornwall

Sample Coordinates

S1 50⁰ 19.861' N; 4⁰ 54.248' W

S2 50⁰ 19.863' N; 4⁰ 54.257' W

S3 50⁰ 19.856' N; 4⁰ 54.248' W

S4 50⁰ 19.809' N; 4⁰ 54.274' W

S5 50⁰ 19.810' N; 4⁰ 54.272' W

S6 50⁰ 19.809' N; 4⁰ 54.275' W

S7 50⁰ 20.014' N; 4⁰ 54.330' W

S8 50⁰ 20.059' N; 4⁰ 54.344' W

S9 50⁰ 20.103' N; 4⁰ 54.321' W

S10 50⁰ 19.559' N; 4⁰ 54.259' W

S11 50⁰ 19.559' N; 4⁰ 54.260' W

S12 50⁰ 19.558' N; 4⁰ 54.260' W

S13 50⁰ 19.547' N; 4⁰ 54.233' W

S14 50⁰ 19.550' N; 4⁰ 54.233' W

S15 50⁰ 19.014' N; 4⁰ 54.330' W

S16 50⁰ 19.359' N; 4⁰ 54.019' W

S17 50⁰ 19.371' N; 4⁰ 54.010' W

S18 50⁰ 19.422' N; 4⁰ 54.097' W

S19 50⁰ 19.701' N; 4⁰ 54.286' W

S20 50⁰ 20.138' N; 4⁰ 54.352' W

166

A2 Sediment analysis

A2.1 Loss on ignition

In soil and sediment geochemistry, organic matter (OM) content can be estimated from

loss on ignition (Sutherland, 1998). The method is simple, rapid and generally in good

agreement with organic carbon (OC) calculated from the dry combustion analyser. A

porcelain crucible was ignited at 550⁰C for 30 minutes in a muffle furnace, then allowed

to cool in a desiccator and accurately weighed. From the bulk dry sediments, 1.0-2.0 g

was placed in the crucible and weighed accurately, then transferred to a muffle furnace

and heated to 550⁰C for 5 hours. The hot crucible, containing the residue, was placed in

the desiccator and cooled to ambient temperature. The crucible containing the ashed

sediment was weighed accurately. The loss on ignition was calculated as a percentage

using the following equation:

where

Ma is the mass of the empty pre-ignited crucible in g

Mb is the mass of the crucible containing the dry sediment in g

Mc is the mass of the crucible containing the ashed sediments in g

A2.2 Radiometric techniques

Radiometric techniques are used to measure the radiation associated with the nuclear

transformations of unstable elements. In general, these methods are based on the

interaction of radiation with detector materials. However, due to the differing properties

of different ionising radiations, various types of detector are employed. For instance,

167

semi-conductor detectors are commonly used for alpha and gamma measurement and,

although the detector materials and configurations are different, in both cases detect the

radiation through ionisation caused. The charges produced in an ionisation event are

collected by the applied voltage (bias) to produce a current. These current pulses are

shaped and magnified using amplifiers, converted from analogue to digital form, sorted

in a multi-channel analyser and, finally, displayed as a spectrum (a plot of energy (x-

axis) against intensity (y-axis). Two types of semi-conductor detectors were employed

throughout this study, Passivated Implanted Silicon (PIPS) Detectors for alpha emitter

measurement and High-Purity Germanium Detectors (HPGe) for gamma emitter

measurement.

Three main factors are crucial when using alpha spectrometry: detector efficiency,

detector resolution and source preparation. Detector efficiency is the proportion of

radiation detected out of the total emitted by the source. Most alpha spectrometry

detectors have efficiency in the range of 25 - 35%. A detector’s energy resolution is the

ability of the detector to discriminate between two signals which are close in energy.

Practically, alpha resolution is calculated from the ratio between the full width at half

maximum (FWHM) and the energy at the centre of the peak. The FWHM is the width

of the peak at 50% of the height of the highest single channel. Generally, alpha

spectrometry resolution is in the range of 40-100 keV, depending on the efficiency of

the chemical separation and quality of alpha source preparation. Source preparation for

alpha counting has different methods; among them, electrodeposition and co-

precipitation are most popular.

A2.2.1 Alpha spectrometry

Alpha measurement requires radiochemical separation of the element of interest from

the matrix, followed by source preparation. The separation begins by spiking the sample

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with a ‘tracer’, usually a known amount of an isotope of the element of interest. This

enables the chemical yield to be calculated according to the equation:

The concentration of an unknown radionuclide in the prepared source can also be

calculated using the following equation:

Several factors need to be taken into consideration when using alpha spectrometry. The

key factors are:

i) Appropriate chemical separation of the radionuclide of interest is needed.

ii) Good alpha source preparation is desirable. The prepared source should be as

thin as possible with a near-weightless amount of the analyte and minimal

deposition of other elements, to avoid peak tailing and obtain good

resolution.

iii) The source should be uniform to ensure reproducible geometry.

iv) The obtained counts should be enough (about 1000 counts) to give an

adequately low counting error.

A2.2.1.1 Sample preparation

For radium, thorium and uranium analysis, radiochemical separation was performed to

prepare alpha sources. These methods for sediment began by totally dissolving or

leaching the solid, before conducting the chemical separation. Three different chemical

methods were employed in this research: acid leaching, total dissolution using

microwave digestion and sequential extraction. Each of these methods aimed to dissolve

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a specific fraction containing the radionuclides of interest. Acid leaching was applied to

release the environmentally-available fraction from the sediment (Moliner-Martinez et

al., 2004). A low pressure closed vessel microwave system was used to obtain the total

concentration of the radionuclides in the sediments. Sequential chemical extraction is a

useful method to study the speciation of radionuclides in sediments because it provides

some insight into geochemical association (Schultz et al., 1998a). The following

sections provide details about sample dissolution and the separation methods employed

for 226

Ra, 230

Th, 234

U and 238

U in sediments.

A2.2.1.2 Acid leaching

A known amount of the sediment (from 0.5 – 2.0 g) was ashed in a muffle furnace at

550⁰ C for 5 hours and then leached with 15.0 mL aqua regia (concentrated

hydrochloric and nitric acids in a 3:1 ratio) at near boiling point for 3 hours. The aim

was to extract the labile fraction of the radionuclides, leaving the fractions associated

with primary minerals. In particular, the leached fractions include those associated with

organic matter and adsorbed onto the surfaces of minerals and secondary phases

(Marsden et al., 2001; Pekala et al., 2010). After leaching, the volume was made up to

50 mL using 0.1 M HNO3.

A2.2.1.3 Total dissolution

During the last decade, microwave digestion was considered an important

radiochemical tool, especially when dealing with naturally-occurring radionuclides

(Michel et al., 2008). For total dissolution of the sediments, 0.2 g of the ashed sediment

was placed in a closed vessel and wetted overnight with a mixture of 1.0 mL deionised

water, 3.0 mL concentrated nitric acid and 6.0 mL concentrated hydrofluoric acid. The

sample was then digested in a microwave oven with ramping time 10 minutes to 140 ⁰C

(~150 psi) and 50 minutes holding time, and this was repeated three times before

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evaporation. Finally, 2.0 mL of 20 % nitric acid was added to the residue and the

volume was made up to 20 mL with deionised water.

A2.2.1.4 Sequential chemical extraction

Sequential extraction is an analytical process that chemically leaches metals out of

specific operationally defined fractions of the solid samples (e.g. soil, sediment and

sludge). The purpose of sequential extraction is, under various environmental

conditions, to release metals selectively from specific, operationally-defined fractions of

the solid into solution. This multi-step procedure (Table A3) intends that all the metals

of concern are extracted from the sample with increasingly aggressive chemical

treatment. The resulting extracts from the different steps are used to determine the

elements’ concentrations. Factors, such as concentration of reagents, reaction

temperature, duration of extraction, agitation and the reaction pH, are critical to

controlling the concentration of metal extracted from the sample. There have been many

proposed methods for trace elements and radionuclides chemical extraction during the

last three decades (Leleyter and Probst, 1999; Outola et al., 2009; Schultz et al., 1998b;

Tessier et al., 1979). The following paragraphs explain in detail the sequential

extraction method used in this study, which is that of Schultz et al. (1998b).

Into a 50.0 mL centrifuge tube, 1.0 g of the dry sediment was weighed and wetted with

deionised water overnight, to help the sediment become hydrated and encourage

swelling of clay minerals. Then, the sequential extraction method was applied to leach

five fractions; namely, exchangeable; bound to organic matter; associated with

carbonates; associated with manganese/iron oxides; and a residual fraction (Fig. A1).

However, because the interest of this study is to explore the radionuclides’ mobility

rather than determine the total concentrations, aqua regia (3:1 HCl/HNO3) instead of an

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HF/HClO4 mixture, as in the original method, was used to extract the residual fraction.

The reagent/sample ratio was kept constant at 15 mL/g.

Following each extraction step, the samples were centrifuged for 30 minutes at 3000

rpm (about 6500 g) and filtered through a Whatman 541 filter paper (Fig. A2). After

each step, the filtrate was saved for radium, thorium, uranium and stable element

measurements, while the solid was retained for the next step.

Table A3 Sequential extraction steps

Fraction Extractive reagents Temp. ⁰C Shaking time (h)

Exchangeable 0.4 M MgCl2 R. T.* 1

Organic matter 5-6% NaOCl (pH7.5) 96 0.5 x 2

Carbonates 1 M NaOAc in 25% HOAc (pH 4) R. T. 2 x 2

Oxides (Fe/Mn) 0.04 M NH2OH.HCl (pH 2) R. T. 5

Residual HCl/HNO3 (3:1) 96 2

*R.T. = Room temperature

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Figure A1 A single grain identifying possible fractions released in sequential extraction

(Kaplan and Serkiz, 2001)

173

Figure A2 Schematic diagram illustrating one separation step in sequential extraction

(Schultz et al., 1998b)

A2.2.1.5 Ra separation

Radiochemical separation of radium was modified from that of (Smith and Mercer,

1970) using 225

Ra (150 mBq) as a radiotracer. Radium was co-precipitated with PbSO4

by adding consecutively 1.0 mL of concentrated H2SO4, 2.0 g K2SO4 and 1.0 ml of 0.24

M of Pb(NO3)2. The solid was centrifuged in a 50.0 mL tube at 3000 rpm (about 6200

g) for 10.0 minutes, and then washed with 20.0 mL of a mixture of 0.2 M H2SO4/0.1 M

K2SO4.

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The precipitate was dissolved in 5.0 mL of 0.1 M ethylenediaminetetraacetic acid


X8, 100-200 mesh, chloride form, 5 x 0.5 cm) to remove sulphate and washed with 13.0

mL 0.01 EDTA/ NH4OH. To the eluate, 1.0 ml 5.0 M CH3COONH4 was added (pH 4.5)

and the solution was passed through a cation exchange column (Bio-Rad AG50W-X12,

200- 400 mesh, 8.0 x 0.7 cm) at a flow rate of 1.0 mL/minute. The column was

previously conditioned with 15.0 mL 1.5 M CH3COONH4 followed by 15.0 mL 0.25 M

CH3COONH4. Another 50.0 mL 1.5 M CH3COONH4/0.1 M HNO3 was passed through

this column to remove Pb and Ac, while Ba was eluted by washing the column with

40.0 mL 2.5 M HCl. Finally, Ra was eluted with 25.0 mL 6.0 M HNO3, and this

solution was evaporated to dryness using a heating lamp.

The electrolysis cell consists of two glass tubes (Sovril SV 30) joined with a SV 30

plastic joint. A polished stainless steel planchette (cathode) was held between the two

glass tubes by a recessed brass planchette mount supported by the lower electrode. The

cell was sealed with a Teflon ring and checked for leaking. A platinum wire anode,

inserted in a narrow glass tube, was passed through a rubber bung into the electrolyte

solution to complete the electric circuit.

For radium electroplating, the Ra fraction was re-dissolved in organic electrolyte

solution (9.0 mL ethanol in 1.0 mL 0.1 M HNO3 in and 2 mL 0.05 HCl ) and

electroplated on to a stainless steel planchette at 120 mA for 30 minutes. One minute

before the end, 1.0 ml of ammonia solution (s.g. 0.88) was added and the power was

turned off. The solution was discarded and the planchette was dried on a hotplate at 200

°C for 5 minutes.

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A2.2.1.6 Th and U separation

The uranium and thorium separations were based on extraction chromatography

methods (Carter et al., 1999; Eichrom Technologies 2001). For acid leaching,

approximately 1.0 g of the ashed sediment was placed in a 150.0 mL glass beaker, and

232U and

229Th tracers (40 mBq and 50 mBq respectively) were added. Following

leaching, as described for Ra, the solution was made up to 50.0 mL with deionised

water, filtered using a Buchner funnel and taken to near dryness under a heating lamp.

For total dissolution, the sample was made up in 20 % HNO3 before evaporation. In

both cases, 5.0 mL conc. HNO3 was added to the residue and the solution brought to

near dryness under a heating lamp. The residue was dissolved with 10.0 mL of 3.0 M

HNO3/1 M Al(NO3)3 and the resultant solution was centrifuged at 3000 rpm (about

6500 g) for 10 minutes.

An extraction chromatography column (TEVA, 2.0 mL pre-packed column; Eichrom

resin, Triskem, France) was preconditioned with 5 ml 3.0 M HNO3 before loading the

leached solution. The beaker was washed with 5.0 mL 3.0 M HNO3 and the wash was

passed through the column. The uranium fraction was collected in a beaker while

thorium was retained on the column. The column was rinsed with 30.0 mL 3.0 M HNO3

which was discarded. Thorium was eluted from the TEVA column with the consecutive

addition of 20.0 mL of 9.0 M HCl and 5.0 mL of 6.0 M HCl. The eluate was evaporated

to near-dryness for electrodeposition in the presence of 1.0 mL 10% (w/v) KHSO4 using

a heat lamp.

To purify the uranium-containing fraction, the eluate was passed through a UTEVA (2.0

mL pre-packed column; Eichrom resin, Triskem, France) separation column, previously

conditioned with 5.0 mL of 3.0 M HNO3. The beaker containing the uranium fraction

was rinsed with 5.0 mL of 3.0 M HNO3 and the washings passed through the column.

176

The column was rinsed with consecutive additions of 5.0 mL of 3.0 M HNO3, 5.0 mL of

9.0 M HCl and 20.0 mL of 5.0 M HCl in 0.05 H2C2O4. All these eluates were discarded

and, finally, uranium was stripped with 15.0 mL of 1.0 M HCl. The eluent was

evaporated to near dryness for electrodeposition in the presence of 1.0 mL 10% (w/v)

KHSO4 using a heat lamp.

For Th/U electrodeposition, 2.5 ml of 5 wt. % NaHSO4, 2.0 ml of deionised water and

5.0 ml of 15.0 wt. % Na2SO4were added to the residue of the purified Th/U fractions

and heated gently until the residue dissolved. The solution was transferred to an

electrodeposition cell and rinsed in with 3.0 ml deionised water, then 1.0 ml of 20.0 g/L

(NH4)2C2O4 plating solution was added. The current was adjusted to 0.5 A for 5 minutes

and then to 0.75 A for 90 minutes. One minute before the end, 2.0 ml of 25.0 wt. %

KOH was added and the power was turned off. The solution was discarded and the cell

was washed with 2.0 ml 5.0 wt. % NH4OH.

Finally, the stainless-steel disk was rinsed consecutively with a small volume of 5.0 wt.

% ammonium hydroxide, ethanol and acetone before being dried on a hotplate at 200 °C

for 5 minutes.

The whole radiochemical separation method was validated using blank samples

(deionised water spiked with the tracer) and standard additions (where a known amount

of 238

U was added to three duplicate samples and then the separation was performed on

the two samples).The blank analyses always gave less than 5 counts in each uranium

region of interest, whereas all the sample analyses are based on signals of at least 100

counts. In the standard additions, the measured uranium recoveries were 89%, 116%

and 87% of the added uranium.

177

A2.2.2 Gamma spectrometry

High-purity germanium (HPGe) detectors are used in gamma spectrometry. The

operational condition of the instrument requires the detector to be cooled to liquid

nitrogen temperature, 77 K, to reduce thermal excitation of electrons and to prevent

damage to the detector. Gamma rays emitted by radionuclides interact with a solid state

detector principally through Compton scattering, pair production and the photoelectric

effect. Among these, the photoelectric effect is the most significant in gamma

spectrometry since the gamma rays lose all of their energy in one interaction. The

spectrum spans a wide range of photon energies, in principle allowing simultaneous

measurement of many radionuclides in the sample. As with alpha spectrometry, in order

to get the activity concentrations of the different radionuclides, a standard is required for

detector efficiency calculation.


Sample preparation for gamma measurement is relatively straightforward. To avoid the

escape of 222

Rn gas, the sample was sealed in a double polypropylene container and put

aside for four weeks. During this time, secular equilibrium between 226

Ra, 222

Rn, 214

Bi

and 214

Pb will be reached and the activity concentrations of all these radionuclides will

be equal. The samples were counted for 12 hours, and the activity concentration of 226

Ra

was estimated from measurements of the 214

Bi gamma line at 609 keV and the 214

Pb

gamma line at 352 keV.

Many considerations in gamma measurements should be taken into account; the crucial

factor being the geometry. The sample and the standard should have the same geometry,

in order to make the calculation simple. The samples were prepared in the same

physical geometries (height, volume and density) as the standard. Since the standard

was prepared by adding a known amount of 226

Ra to two samples, one with low organic

178

content and one with high organic content, which both had low radium contents, this

allows compensation for the effects of chemical composition.

A2.3 Mineral analysis techniques

A2.3.1 Powder X-ray diffraction (XRD)

X-ray powder diffraction (XRD) is a common technique for studying the structure of

crystals and atomic spacing of minerals. The technique depends on developing an

interference pattern through interaction of a monochromatic X-ray beam with the

crystals of the sample. Mineral identifications are made by comparing the pattern

obtained with a database. The cathode ray tube generates the X-ray by heating a

filament to produce electrons. These electrons are accelerated by applying voltage,

before hitting the target to generate characteristic X-ray spectra which are then filtered

to produce monochromatic X-rays. The detector records and processes the scattered X-

ray signal.


The dry sediments were sieved through 80 mesh and a suitable amount (~0.5 g) was

placed on the sample holder. A smooth, flat surface was obtained using a glass slide,

before placing the sample in the specimen position of the XRD. Mineral identifications

were made using a Bruker D8Advance Powder diffractometer. The X-ray is generated

from a Cu X-ray tube (Kα with a wavelength of 0.152 nm, current 30 mA at 40 kV) and

the instrument is equipped with a standard scintillation detector. The scanning angle for

the specimen was set from 5⁰ to 75⁰ with a step size of 0.02⁰/s and an exposure time of

30 minutes. Phase identification was performed using Eva 14, Bruker version 2008

analysis tool.

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A2.3.2 Scanning electron microscopy (SEM)

Scanning electron microscopy (SEM) is a powerful microscopic technique for sediment

and particle characterisation. It offers semi-quantitative analysis of the particle surface,

as well as information on morphology. Samples can be characterised using secondary

electron imaging (SEI), backscatter electron imaging (BSE) and energy dispersive X-

ray spectrometry (EDX). Secondary electrons have very low energy (<50 keV) and are

emitted near the surface of the object so the SEI image is valuable in showing

morphology and topography, particularly of sample coatings. Compared with secondary

electrons, BSE have higher energy and are produced over a greater depth profile. Most

importantly, BSE production is atomic number dependent so, the brighter the image, the

higher the atomic number of the element. Accordingly, the BSE image provides

information about the composition of the object. Energy dispersive X-ray spectroscopy

(EDX/EDS) is used to estimate the chemical composition of materials down to a spot

size of a few microns and, in this study, was used to obtain qualitative and semi-

quantitative chemical compositions of selected spots.

A2.3.2.1 Heavy liquid separation

The heavy liquid separation technique was used to separate minerals in solid samples,

based on density to heavy minerals and light minerals. The sample was placed into a

50.0 mL centrifuge tube and a heavy liquid for density separation, LST Fastfloat

consists of sodium heteropolytungstates dissolved in water with a density 2.8 ± 0.02

g/mL, was poured to half-fill the tube. The tube was hand-shaken to mix the grains with

the heavy liquid. Then, the tube was filled with more LST until almost full, and left

overnight to allow the minerals to separate. Once the minerals had separated, the lower

end of the centrifuge tube was immersed into a small container of liquid nitrogen until

the bottom 1 cm of liquid was frozen. The unfrozen solution was decanted and filtered

under gravity and deionised water was used carefully to rinse out any minerals

180

remaining in the tube, while avoiding melting the frozen layer. The bottom layer was

then allowed to melt, and filtered under gravity, rinsing with deionised water. The

filtered samples were rinsed 4-5 times with deionised water to ensure removal of LST,

and the filter papers were placed overnight inside an oven to dry at 100 °C.


Four sediment samples (S3, S7, S13 and S20) were selected for characterisation by the

JEOL JSM-6400 SEM. For the two samples, with the highest radioactivity, S3 and S7,

three subsamples (total sample, light minerals and heavy minerals) were prepared. Each

dry sample was embedded on a glass slide using epoxy resin, and polished to provide a

homogeneous surface for analysis. The samples were carbon-coated so the samples

were conductive, to prevent charging of the surface and to promote emission of

secondary electrons. At the beginning of the analysis, backscattered images were

obtained to localise heavy elements. This was followed by obtaining secondary electron

images from the near surface of the most interesting spots using a voltage of 15-20 kV.

In addition, the EDX Princeton Gamma Tech EDS system was used to perform

elemental analysis.

A2.3.3 Electron microprobe analyser (EMPA)

An Electron Micro Probe Analyser (EMPA) was used to characterise solid samples

using BSE images and wavelength dispersive spectroscopy (WDS) images. BSE images

offer high-resolution maps related to sample composition for distinguishing different

phases. It facilitates location of areas of interest, in particular high atomic number

elements (e.g. U-rich spots). Energy dispersive spectroscopic analysis provides a

quantitative analysis of elemental compositions. WDS analysis provides X-ray maps

over an area of approximately 2.5 µm2 and a depth of 0.5 µm.

181


The same four sediment samples (S3, S7, S13 and S20) that were characterised by

SEM, were further analysed using the CAMECA SX100 EMPA. For the two samples

with the highest radioactivity, S3 and S7, three subsamples (total sample, light minerals

and heavy minerals) were prepared, as for the SEM study. Each dry sample was

embedded on a glass slide using epoxy resin, then polished and carbon-coated to be

conductive. Firstly, a backscattered image of a size of 100 x 100 µm of the sample was

obtained to locate higher atomic number elements. This was followed by selecting a

single grain (typically 50 x 50 µm) for elemental mapping. In addition, the elemental

composition (expressed as the oxides) of 21 major and trace elements was determined

by energy-dispersive spectroscopy at selected spots. During analysis, the acceleration

voltage was 15 kV and the beam current of the probe was 20 nA, and several standards

were used for calibration.

The instrument is equipped with five wavelength detectors and it was also possible to

use these to obtain elemental maps using wavelength-dispersive spectroscopy. The 10

elements selected were divided into two groups. The first group included U, Ca, Mg,

Mn and Fe, while K, Cu, As, Sn and Pb were in the second group.

A2.4 Inductively coupled plasma mass spectroscopy (ICP-MS)

Since the development of ICP-MS in the 1980s, the technique has become widely used

to determine trace element concentrations and isotopic ratios in environmental

materials. ICP-MS detects trace and ultra-trace concentrations (ppm-ppt)

simultaneously with high accuracy and precision, within a short analysis time and,

recently, measurements of a range of long-lived radionuclides in the environment have

been reported (Becker, 2005; Hou and Roos, 2008; Lariviere et al., 2006).

182

The principle of ICP-MS involves three main steps. Firstly, the chemical species in the

sample are decomposed to their atomic constituents and ionised in inductively coupled

argon plasma (ICP). This ICP is characterised by extremely high temperatures (6000-

8000 K), which are sufficient to produce a high degree of ionisation (> 90 % for most

elements) with a low percentage (~ 1%) of multiple charged ions. The second step

transfers the positive ions from the ICP, formed at atmospheric pressure, to the high

vacuum of the mass spectrometer (MS) via an interface. The final step separates the

positive ions by mass, according to their mass/charge ratio, followed by measurement in

an ion detector.

The aqueous samples in this study were introduced to the ICP-MS as mists, using a

nebuliser. The mist contains two components; droplets with a larger size and aerosols

with small size. The latter passed through a chamber, where the bigger droplets are

collected at the bottom of the chamber, and discarded. The smaller aerosol droplets are

carried through a hole and mixed with large volume of argon gas, and all travel toward

the torch. The torch is powered by a radio frequency circuit resulting from a coil of

hollow copper wires surrounded by a silicon tube. The torch operates at extremely high

temperatures (around 104 K) to maintain dissociation, atomisation and ionisation of the

aerosol in the plasma. The output of the plasma is then introduced to an interface

comprised of two nickel cones; sampler and skimmer. The aim of the interface is to

allow the ions (generated at atmospheric pressure) to pass into the quadrupole mass

spectrometer (operating under a vacuum) and eliminate atoms and uncharged species.

The quadrupole mass analyser consists of two pairs of rods aligned in a parallel pattern.

A direct current (DC) is applied to one pair, and a radio frequency (RF) is placed on the

other. This allows an ion with a selected mass/charge ratio (m/z) to reach the detector,

183

while any unrequired ions are removed from the detector. Sweeping of the combination

between DC and RF allows different m/z ratio ions to be measured in the detector.

The ions are counted in a dynode electron multiplier detector. When an ion coming out

from the quadrupole hits the inner surface of the first dynode, secondary electrons are

released. The secondary electrons strike the next dynode and release more electrons.

The process continues to produce electric pulses, which are counted by the integrated

circuit. The magnitude of the electrical pulses corresponds to the concentration of

analyte in the sample. Quantitative analysis of samples requires comparing the signal

from the analyte with the signal of a matrix-matched standard containing a known

concentration of the same analyte. In practice, a series of standard solutions containing

concentrations spanning the range of expected analyte concentrations is used to produce

a calibration curve.

A2.4.1 Sample preparation

Two types of sample were analysed by ICP-MS; sequential extraction leachates from

sediments and water samples from Cornwall. For trace element analysis of sediment

samples, 1 ml of each fraction obtained from the sequential extraction was made up to

10 mL with 2 % nitric acid. The samples were run in the ICP-MS using a standard

solution prepared for each analyte. Water samples for cation measurements were filtered

and acidified in the field with nitric acid to a pH < 2, then analysed directly.

184

Figure A3 Basic components of an ICP-MS. Adapted from Thomas (2008)

A3 Water analysis

A3.1 Physicochemical properties

A3.1.1 pH

A pH meter, SevenEasy, Mettler-Toledo GmbH, and a probe were used to measure the

pH of the water samples from the River Fal in Cornwall. Prior to the measurement, the

pH meter was calibrated using three buffers with pH 4, pH 7 and pH 10. In a 50 mL

glass beaker, sufficient amounts of the sample, to immerse the probe, were placed. The

probe was rinsed with deionised water and a small amount of the sample. Then, the

probe was immersed in the sample, and the pH was taken after the reading stabilised.

A3.2.2 Electrical conductivity (EC)

Electrical conductivity (EC) is a measure of the total dissolved ionic species in an

aqueous phase and can be measured using a meter and a probe. Specific conductivity is

expressed as Siemens per centimetre (S/cm). A Jenway 4010 conductivity meter was

185

calibrated before sample analysis using 0.1 M KCl standard solution. Sufficient amount

of the sample, to immerse the probe, was placed in a 50 ml glass beaker. The probe was

rinsed with deionised water and a small amount of the sample. Then, the probe was

immersed in the sample and the EC reading was taken after it had stabilised.

A3.2 Ion chromatography

In the mid-1970s, ion chromatography (IC) was introduced in water chemistry, to

determine cations and anions. Currently, it is a common analytical method for anions

including nitrate, sulphate, fluoride and chloride (Ohta et al., 2000). IC is based on

separation of ionic species through their interactions, in the mobile phase, with the

resin. Following separation, the concentrations of the ions are measured. The separation

is based on size and/or affinity for the stationary phase. An IC instrument comprises

three main parts: a stationary phase of low ion-exchange capacity resin; a detector; and

a suppressor column to improve the separation and detection sensitivity (López-Ruiz,

2000). Briefly, when the aqueous phase passes through the pressurised column, ions are

adsorbed. The eluent passes through the column to release the adsorbed species with

differing retention times. IC has many advantages: the procedure is rapid and simple, it

can be used to distinguish between oxo-ions (e.g. nitrate/nitrite) and it requires only a

small quantity of the sample for analysis (Woods and Rowland, 1997).

A3.2.1 Sample preparation

All reagents were of analytical reagent grade, and deionised Milli-Q water was used in

all preparation and measurement. The filtered water samples were measured in the IC

for fluoride, chloride, bromide, nitrate, nitrite, phosphate and sulphate. The IC

instrument consisted of a Metrohm 761 Compact ion chromatograph, fitted with a

Dionex Ion-Pac AG9-HC (guard), a Dionex Ion-Pac AS9-HC analytical column and a

conductivity detector. The backpressure was 2000 psi, the mobile phase was 9 mM

186

Na2CO3 and the eluent flow rate was 1.4 mL/min. A set of standard solutions, with

concentrations of 0.5, 3.0, 10.0 and 30.0 mg/L for chloride, nitrate and sulphate; 0.1,

0.5, 1.0 and 3.0 mg/L for bromide, nitrite and phosphate, was used for calibration. The

detection limit was approximately 0.05 mg/L for most analytes.

187

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